PCBs Cancer Dose Response Assessment and Application to Environmental Mixtures, September 1996

EPA/600/P–96/001F

September 1996

PCBs: Cancer Dose-Response Assessment

and Application to Environmental Mixtures

National Center for Environmental Assessment

Office of Research and Development

U.S. Environmental Protection Agency

Washington, DC

DISCLAIMER

This document has been reviewed in accordance with U.S. Environmental

Protection Agency policy and approved for publication. Mention of trade names or

commercial products does not constitute endorsement or recommendation for use.

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CONTENTS

LIST OF TABLES .................................................... v

PREFACE ........................................................... vi

AUTHORS, CONTRIBUTORS, AND REVIEWERS .......................... viii

1. INTRODUCTION .................................................. 1

1.1. PCB MIXTURES ........................................... 1

1.2. CANCER POTENTIAL OF PCB MIXTURES ...................... 5

1.3. APPROACH TAKEN BY THIS ASSESSMENT .................... 6

2. SUMMARY OF STUDIES USED IN THE DOSE-RESPONSE ASSESSMENT . . . 8

2.1. CANCER STUDIES IN HUMANS .............................. 8

2.2. LIFETIME CANCER STUDIES IN ANIMALS ...................... 10

2.3. PARTIAL LIFETIME STUDIES IN ANIMALS ...................... 17

2.4. TUMOR INITIATING AND PROMOTING ACTIVITY ................ 19

2.5. ABSORPTION AND RETENTION .............................. 22

2.6. METABOLISM AND MODE OF ACTION IN THE LIVER ............. 25

2.7. MODE OF ACTION IN THE THYROID .......................... 26

3. DOSE-RESPONSE ASSESSMENT ................................... 28

3.1. APPROACHES TO DOSE-RESPONSE ASSESSMENT ............. 28

3.2. BIOLOGICALLY BASED MODELING OF TUMOR PROMOTION ...... 31

3.3. EMPIRICAL MODELING OF TUMOR INCIDENCE ................. 32

3.4. ANALYSES OF CONGENER TOXICITY ......................... 37

4. APPLICATION OF THE DOSE-RESPONSE ASSESSMENT ................ 39

4.1. APPLICATION TO PCB MIXTURES IN THE ENVIRONMENT ........ 39

4.2. APPLICATION TO DIFFERENT ROUTES OF EXPOSURE .......... 45

4.3. APPLICATION TO LESS-THAN-LIFETIME AND EARLY-LIFE

EXPOSURES ............................................ 46

4.4. APPLICATION WITH DIOXIN TOXIC EQUIVALENCE FACTORS ..... 49

5. CHARACTERIZATION AND GUIDANCE FOR RISK ASSESSORS ........... 50

5.1. DOSE-RESPONSE CHARACTERIZATION ....................... 50

5.2. INFLUENCE OF PROPOSED CANCER GUIDELINES .............. 55

5.3. RESEARCH NEEDS ........................................ 56

5.4. SUMMARY OF GUIDANCE FOR RISK ASSESSORS .............. 59

5.5. EXAMPLES ............................................... 61

6. REFERENCES ................................................... 65

APPENDIX: EMPIRICAL MODELING RESULTS ........................... 73

LIST OF TABLES

1–1 Typical composition (%) of some commercial PCB mixtures ............... 2

1–2 Domestic production (%) of commercial PCB mixtures, 1957–1977 ......... 2

2–1 Liver tumor incidences in rats from lifetime exposure studies, 1975–1985 . . . 11

2–2 Liver tumor incidences in rats from 1996 lifetime exposure study .......... 14

2–3 Liver tumor incidences in female rats from 1996 stop study .............. 15

2–4 Thyroid gland tumor incidences in male rats from 1996 lifetime

exposure study ................................................ 16

2–5 Mixtures and congeners tested for tumor promoting activity .............. 21

3–1 Human potency and slope estimates derived from rat liver tumors ......... 34

3–2 Range of human potency and slope estimates ........................ 35

3–3 PCB congeners of highest concern ................................. 37

3–4 WHO interim TEFs for human intake of dioxin-like PCBs ................ 38

4–1 Tiers of human potency and slope estimates for environmental mixtures .... 44

5–1 Sample lifetime average daily dose calculations (examples 1 and 2) ....... 62

5–2 Sample risk calculations (example 1) ............................... 62

5–3 Sample risk calculations (example 2) ............................... 63

5–4 Sample congener concentrations and dioxin toxic equivalents (TEQs)

in edible portion of fish (example 3) ................................. 64

5–5 Sample lifetime average daily dose calculations (example 3) ............. 64

5–6 Sample risk calculations (example 3) ............................... 65

PREFACE

This report updates the cancer dose-response assessment for PCBs and shows

how information on toxicity, disposition, and environmental processes can be

considered together to evaluate health risks from PCB mixtures in the environment.

Intended to be brief, it focuses on analysis and interpretation rather than a compilation

of study results. More detailed information on PCB toxicity has been compiled by the

Agency for Toxic Substances and Disease Registry (ATSDR, 1993, 1995), Safe (1994),

Silberhorn et al. (1990), and the U.S. Environmental Protection Agency (U.S.EPA)

(1988a).

Although not covered by this report, PCBs also have significant ecological and

human health effects other than cancer, including neurotoxicity, reproductive and

developmental toxicity, immune system suppression, liver damage, skin irritation, and

endocrine disruption. Toxic effects have been observed from acute and chronic

exposures to PCB mixtures with varying chlorine content. These toxic effects should be

included along with cancer in future assessments of PCBs.

This report is to be used to support risk-based decisions within the general

policy framework provided by applicable EPA statutes and does not alter such policies.

It does not imply that one kind of information or another is a prerequisite for action. Not

every risk assessment based on this dose-response assessment will have the same

scope or depth; the level of detail of an assessment is a matter of management policy.

This report is being made available to the public and the U.S. Congress,

responding to the report of the House of Representatives Appropriations Committee,

which specifies:

vii

By December 31, 1995, the Administrator shall submit to the Congress,

and make available to the public, a draft report providing an assessment

of the risk of each of the polychlorinated biphenyl (PCB) mixtures that has

been the subject of laboratory animal cancer bioassays, and a proposed

methodology for assigning cancer risk numbers to mixtures of PCB's

found in the environment. By September 1, 1996, the Committee directs

that EPA shall have completed, by a panel of independent experts on the

carcinogenicity of PCB's, a peer review of the draft report, and shall

submit a final report to the Congress and make it available to the public.

A new laboratory animal study of four commercial mixtures will soon be made

public. Because this study will provide the most comprehensive information for dose-

response modeling, this report makes use of preliminary information obtained through

July 1996. Although some of this information is still under review and additional

information may soon become available, updating the dose-response assessment at

this time allows current decisions to reflect current science and provides a framework

for incorporating new information.

viii

AUTHORS, CONTRIBUTORS, AND REVIEWERS

This report was written by Dr. Jim Cogliano of EPA's National Center for

Environmental Assessment. The author gratefully acknowledges the helpful and

insightful comments from many scientists in EPA's program, regional, and research

organizations:

Air and Radiation: Jane Caldwell.

Prevention, Pesticides and Toxic Substances: David Lai,

Elizabeth Margosches.

Solid Waste and Emergency Response: Dorothy Canter.

Water: Robert Cantilli, Michael Cox, Krishan Khanna, Edward Ohanian.

Region 1 (Boston): Mary Ballew, Andy Beliveau, Ann-Marie Burke,

Jui-Yu Hsieh, Ronnie Levin, Margaret McDonough, Marybeth Smuts.

Region 2 (New York): Marian Olsen, Anita Street, Doug Tomchuk.

Region 3 (Philadelphia): Debra Forman.

Region 5 (Chicago): Carole Braverman, Milton Clark, Stephen Johnson.

Region 6 (Dallas): Young-Moo Kim, Maria Martinez, Jeffrey Yurk.

Region 7 (Kansas City): Dave Crawford.

Region 8 (Denver): Bob Benson, Suzanne Wuerthele.

Region 9 (San Francisco): Arnold Den, Daniel Stralka.

Region 10 (Seattle): Dana Davoli, Roseanne Lorenzana.

Research and Development: Linda Birnbaum, Ruth Bleyler, William Farland,

Charli Hiremath, James Holder, Prasad Kodavanti, Aparna Koppikar,

Jim Lake, Susan Norton, Charles Ris, Cheryl Siegel Scott, Dharm Singh,

Jeanette Wiltse.

The author also gratefully acknowledges Dr. Stephen Hamilton of the General

Electric Company for providing preliminary results of the new cancer studies.

A draft report was made available to the public in January 1996. The following

people or organizations provided written comments:

Obaid Faroon, Agency for Toxic Substances and Disease Registry, Atlanta GA.

Richard W. Green, Delaware Department of Natural Resources and

Environmental Control, Dover DE.

Marion Harnois, Massachusetts Department of Environmental Protection,

Boston MA.

Greg Schweer, Versar Inc., Springfield VA.

Tracey Slayton, Gradient Corporation, Cambridge MA.

Chemical Land Holdings, Inc., Kearny NJ.

The draft report was considered at a public, external peer review workshop in

May 1996. A workshop report was written by the review panel, which included the

following experts on the carcinogenicity of PCBs:

Lucy Anderson, National Cancer Institute, Frederick MD.

John Brown, Jr., General Electric Company, Schenectady NY.

Peter DeFur, Environmental Stewardship Concepts, Richmond VA.

Dale Hattis, Clark University, Worcester MA.

Kim Hooper, California Environmental Protection Agency, Berkeley CA.

Marty Kanarek, University of Wisconsin-Madison, Madison WI.

Nancy Kim, New York State Department of Health, Albany NY.

Loren Koller (chair), Oregon State University, Corvallis OR.

John Moore, Institute for Evaluating Health Risks, Washington DC.

Christopher Portier, National Institute of Environmental Health Sciences,

Research Triangle Park NC.

Paul Price, McLaren/Hart ChemRisk, Portland ME.

Larry Robertson, University of Kentucky, Lexington KY.

Stephen Safe, Texas A&M University, College Station TX.

Susan Velazquez, Toxicology Excellence for Risk Assessment, Cincinnati OH.

Gary Williams, American Health Foundation, Valhalla NY.

At the workshop, there was opportunity for public comment. The following

people provided oral comments:

Brent Finley, McLaren/Hart ChemRisk, Alameda CA.

John Schell, TERRA, Inc, Tallahassee FL.

Thomas Starr, ENVIRON Corporation, Arlington VA.

This report has been approved by EPA's consensus review panel for inclusion

on EPA's Integrated Risk Information System (IRIS).

The author is grateful for the contributions of the review panel and all who

commented on the draft report. Their efforts truly improved the final product.

Some notes on chemical structure and nomenclature: Each PCB molecule consists of two 6-carbon rings, with one

chemical bond joining a carbon from each ring (imagine sunglasses with hexagonal frames). Chlorine can attach to any

of the other 10 carbons; these positions are said to be substituted

. There are 209 possible arrangements, called

congeners

; congeners with the same number of chlorines are called isomers. The number and position of chlorines

determine a molecule's physical and chemical properties. The 10 positions are numbered 2–6 on one ring and 2'–6' on

the other. For example, the congener 2,4,2',5'-tetrachlorobiphenyl has chlorines in positions 2 and 4 of one ring and 2

and 5 of the other. (Standard chemical notation for this congener is 2,2',4,5'-tetrachlorobiphenyl; instead, this

assessment lists chlorines on one ring, then the other, to emphasize each ring's chlorination pattern.) Positions 2, 6, 2',

and 6', adjacent to the bond, are called ortho

positions; 3, 5, 3', and 5', meta positions; 4 and 4' (the outermost), para

positions. The International Union of Pure and Applied Chemists (IUPAC) has adopted an alternative system for

numbering congeners sequentially from 1 to 209; numbers assigned to congeners named in this assessment are listed

in table 3–3. A molecule's two rings can twist on the bond joining them; they are coplanar

if aligned in the same plane.

Chlorine in ortho positions inhibits a coplanar alignment. Coplanar molecules have dioxin-like properties (Safe, 1990,

1994; U.S. EPA, 1994b). PCB mixtures manufactured in the United States carried the trademark "Aroclor" followed by a

four-digit number; the first two digits are "12," and the last two digits indicate the percent chlorine content by weight. For

example, Aroclor 1260 contains approximately 60 percent chlorine by weight. Aroclor 1016 is an exception to this

scheme; it contains approximately 41 percent chlorine. "Clophens" and "Kanechlors" are PCB mixtures manufactured in

Germany and Japan, respectively; these series have their own numbering schemes.

1. INTRODUCTION

1.1. PCB MIXTURES

PCBs (polychlorinated biphenyls) are mixtures of synthetic organic chemicals.

Different mixtures can take on forms ranging from oily liquids to waxy solids. Although

their chemical properties vary widely, different mixtures can have many common

components. Table 1–1 shows the overlapping composition of some commercially

manufactured mixtures. Because of their inflammability, chemical stability, and

insulating properties, commercial PCB mixtures had been used in many industrial

applications, especially in capacitors, transformers, and other electrical equipment.

These chemical properties, however, also contribute to the persistence of PCBs after

they are released into the environment. Because of evidence that PCBs persist in the

environment and cause harmful effects, domestic manufacture of commercial mixtures

was stopped in 1977; existing PCBs, however, continue in use. Table 1–2 shows some

commercial mixtures as a percentage of domestic production.

Table 1–1. Typical composition (%) of some commercial PCB mixtures

Aroclor Clophen Kanechlor

1016

1242 1248 1254 1260 A 30 A 60 300 400 500

Mono-CBs 2 1 — — — — — — — —

Di-CBs 19 13 1 — — 20 — 17 3 —

Tri-CBs 57 45 21 1 — 52 — 60 33 5

Tetra-CBs 22 31 49 15 — 22 1 23 44 26

Penta-CBs — 10 27 53 12 3 16 1 16 55

Hexa-CBs — — 2 26 42 1 51 — 5 13

Hepta-CBs — — — 4 38 — 28 — — —

Octa-CBs ———— 7 — 4 ———

Nona-CBs ———— 1 —— ———

Deca-CB ————— —— ———

Columns may not total 100% due to rounding; "—" signifies less than 1%.

Lot-to-lot variability exists but has not been quantified.

Impurities include chlorinated dibenzofurans and naphthalenes; see World Health Organization

(WHO) (1993) for sample concentrations.

Sources: Adapted from Silberhorn et al. (1990), ATSDR (1995).

Table 1–2. Domestic production (%) of commercial PCB mixtures, 1957–1977

Percent of

Mixture

production

Aroclor 1016 13

Aroclor 1221 1

Aroclor 1232 < 1

Aroclor 1242 52

Aroclor 1248 7

Aroclor 1254 16

Aroclor 1260 11

Aroclor 1262 1

Aroclor 1268 < 1

Column does not total 100% due to rounding.

Source: Adapted from Brown (1994).

In the environment, PCBs also occur as mixtures of congeners, but their

composition differs from the commercial mixtures. This is because after release into

the environment, the composition of PCB mixtures changes over time, through

partitioning, chemical transformation, and preferential bioaccumulation.

Partitioning refers to processes by which different fractions of a mixture separate

into air, water, sediment, and soil. PCBs adsorb to organic materials, sediments, and

soils; adsorption tends to increase with chlorine content of the PCBs and organic

content of the other material (Callahan et al., 1979). PCBs can volatilize or disperse as

aerosols, providing an effective means of transport in the environment (Callahan et al.,

1979). Congeners with low chlorine content tend to be more volatile and also more

soluble in water (Callahan et al., 1979). Vaporization rates and water solubility of

different Aroclors and individual congeners vary over several orders of magnitude

(Hutzinger et al., 1974; Erickson, 1986).

Biodegradation transforms the chemical composition of PCB mixtures in the

environment. Anaerobic bacteria in sediments selectively remove chlorines from meta

and para positions, appearing to reduce the toxicity and bioaccumulation potential of

residues; the occurrence and extent of these dechlorinations can be limited by

sediment PCB concentrations (Abramowicz, 1990; Brown and Wagner, 1990; Lake

et al., 1992). (Dechlorination is not synonymous with detoxication, as congeners

having carcinogenic activity can be formed through dechlorination.) Aerobic bacteria

remove chlorines from PCBs with low chlorine content (1–4 chlorines) and break open

the carbon rings through oxidation (Abramowicz, 1990). PCBs with higher chlorine

content are extremely resistant to oxidation and hydrolysis (Callahan et al., 1979).

Photolysis can slowly break down congeners with high chlorine content (Callahan

et al., 1979). Overall, dechlorination processes are slow and altered PCB mixtures can

persist in the environment for many years.

PCBs can accumulate selectively in living organisms. PCBs are highly soluble

in lipids and are absorbed by fish and other animals. Rates of metabolism and

elimination are slow and vary by congener (Matthews and Anderson, 1975).

Bioaccumulation through the food chain tends to concentrate congeners of higher

chlorine content, producing residues that are considerably different from the original

Aroclors (Schwartz et al., 1987; Oliver and Niimi, 1988; Lake et al., 1995). PCB

residues in fish and turtles, changed through environmental or metabolic alteration,

could not be characterized by Aroclor 1242, 1248, 1254, or 1260 standards (Schwartz

et al., 1987). Congener distributions in several species, including humans, do not

resemble any Aroclor (McFarland and Clarke, 1989). Because, in general, some toxic

congeners are preferentially retained, bioaccumulated PCBs appear to be more toxic

than commercial PCBs (Aulerich et al., 1986; Hornshaw et al., 1983).

PCBs are widespread in the environment, and humans are exposed through

multiple pathways. Levels in air, water, sediment, soil, and foods vary over several

orders of magnitude, often depending on proximity to a source of release into the

environment (ATSDR, 1993; WHO, 1993). Average daily intake by humans via

ambient air is about 100 ng, and about an order of magnitude higher if indoor

concentrations are considered (ATSDR, 1993). Average daily intake via drinking water

is less than 200 ng (ATSDR, 1993). Estimates of average daily intake via diet vary

widely depending on geographic area, food habits, and sampling methodology;

5–15 g is considered a good estimate of average daily intake via diet in industrialized

countries (WHO, 1993). For nursing infants, average daily intake was estimated at

1.5–27 g/kg (ATSDR, 1993); another study estimated 3–11 g/kg (WHO, 1993). Using

the narrower range, average daily intake for a 5-kg nursing infant would be 15–55 g,

about triple the average adult intake, and approximately 50-fold higher when adjusted

for body weight. Nursing infants are, therefore, an important potentially highly exposed

population. Another is people whose diet is high in game fish, game animals, or

products of animals contaminated through the food chain.

Although environmental mixtures are often characterized in terms of Aroclors,

this can be both imprecise and inappropriate. Qualitative and quantitative errors can

arise from judgments in interpreting gas chromatography/mass spectrometry (GC/MS),

which reveals a spectrum of peaks that are compared with characteristic patterns for

different Aroclors. For environmentally altered mixtures, an absence of these

characteristic patterns can suggest the absence of Aroclors, even though some

congeners are present in high concentrations. Large differences have been found in

results reported by laboratories analyzing the same sediment samples (Alford-Stevens

et al., 1985; Alford-Stevens, 1986).

1.2. CANCER POTENTIAL OF PCB MIXTURES

Occupational studies show some increases in cancer mortality in workers

exposed to PCBs. Bertazzi et al. (1987) found significant excess cancer mortality at all

sites combined and in the gastrointestinal tract in workers exposed to PCBs containing

54 and 42 percent chlorine. Brown (1987) found significant excess mortality from

cancer of the liver, gall bladder, and biliary tract in capacitor manufacturing workers

exposed to Aroclors 1254, 1242, and 1016. Sinks et al. (1992) found significant excess

malignant melanoma mortality in workers exposed to Aroclors 1242 and 1016. Some

other studies, however, found no increases in cancer mortality attributable to PCB

exposure (ATSDR, 1993). The lack of consistency overall limits the ability to draw

definitive conclusions from these studies. Incidents in Japan and Taiwan where

humans consumed rice oil contaminated with PCBs showed some excesses of liver

cancer, but this has been attributed, at least in part, to heating of the PCBs and rice oil,

causing formation of chlorinated dibenzofurans (ATSDR, 1993; Safe, 1994).

A new study of rats fed diets containing Aroclors 1260, 1254, 1242, or 1016

found statistically significant, dose-related, increased incidences of liver tumors from

each mixture (Brunner et al., 1996). Earlier studies found high, statistically significant

incidences of liver tumors in rats ingesting Aroclor 1260 or Clophen A 60 (Kimbrough

et al., 1975; Norback and Weltman, 1985; Schaeffer et al., 1984). Partial lifetime

studies found precancerous liver lesions in rats and mice ingesting PCB mixtures of

high or low chlorine content.

Several mixtures and congeners test positive for tumor promotion (Silberhorn

et al., 1990). Toxicity of some PCB congeners is correlated with induction of mixed-

function oxidases; some congeners are phenobarbital-type inducers, some are

3-methylcholanthrene-type inducers, and some have mixed inducing properties

(McFarland and Clarke, 1989). The latter two groups most resemble 2,3,7,8-

tetrachlorodibenzo-p-dioxin in structure and toxicity.

Overall, the human studies have been considered to provide limited (IARC,

1987) to inadequate (U.S. EPA, 1988a) evidence of carcinogenicity. The animal

studies, however, have been considered to provide sufficient evidence of

carcinogenicity (IARC, 1987; U.S. EPA, 1988a). Based on these findings, some

commercial PCB mixtures have been characterized as probably carcinogenic to

humans (IARC, 1987; U.S. EPA, 1988a). There has been some controversy about how

this conclusion applies to PCB mixtures found in the environment.

1.3. APPROACH TAKEN BY THIS ASSESSMENT

Previous assessments developed a single dose-response slope (7.7 per

mg/kg-d average lifetime exposure) for evaluating PCB cancer risks (U.S. EPA, 1988a).

With no agreed-on basis for reflecting differences among environmental mixtures, this

slope was used by default for any mixture. Different alternatives have been suggested

that would make some distinctions about cancer risks from different PCB mixtures.

One alternative would assume there is no cancer hazard from environmental mixtures

unless the mixture is highly chlorinated, for example, an overall chlorine content of

approximately 60 percent or greater (Delaware Department of Natural Resources and

Environmental Control, 1994). Another alternative would develop a separate

assessment for each commercial mixture that has been studied. These alternatives

begin to distinguish among PCB mixtures, but they do not address how the

environmental processes of partitioning, transformation, and bioaccumulation diminish

the similarity of environmental mixtures to any of the commercial mixtures.

This new assessment adopts a related approach that distinguishes among PCB

mixtures by using information on environmental processes. Environmental processes

have profound effects that can decrease or increase toxicity, so toxicity of an

environmental mixture is only partly determined by the original commercial mixture.

This new assessment, therefore, considers all cancer studies (which used commercial

mixtures only) to develop a range of dose-response slopes, then uses information on

environmental processes to provide guidance on choosing an appropriate slope for

representative classes of environmental mixtures and different exposure pathways.

Different kinds of information, many not typically considered in dose-response

assessments, are used in this approach. Other innovative features include:

A range of upper-bound potency estimates for PCB mixtures, plus a range

of central estimates, with guidance for choosing estimates from these

ranges to reflect the effect of environmental processes on a mixture's

toxicity. Sources of uncertainty in these estimates are identified and

discussed.

A tiered approach that can use site-specific congener information when

available, but can be adapted if information is limited to total PCBs

encountered through each exposure pathway.

Application of EPA's proposed cancer guidelines (U.S. EPA, 1996a) in the

quantitative dose-response assessment, including the interagency

consensus cross-species scaling factor (U.S. EPA, 1992b) and discussion

of circumstances affecting cancer risk.

A new rat study (Brunner et al., 1996), with parallel experiments for Aroclors

1260, 1254, 1242, and 1016, will soon be made public. Each experiment tested both

sexes at several dose levels. Because this study will provide the most comprehensive

information for dose-response modeling, this assessment makes use of preliminary

information that could be obtained through July 1996. To ensure the scientific quality

of this information, the laboratory report was reviewed by four members of the external

peer review panel (Koller, 1996).

Section 2 briefly summarizes the studies used in developing the dose-response

assessment and applying it to environmental mixtures. For a comprehensive

discussion of PCB toxicity, including many other studies, see ATSDR (1993), Safe

(1994), Silberhorn et al. (1990), or U.S. EPA (1988a). Section 3 uses the studies

summarized in section 2 to develop a new dose-response assessment. Section 4

Standardized mortality ratio (SMR) = 100 × observed / expected.

CI = 95% confidence interval.

discusses application of the dose-response assessment to environmental mixtures, to

different exposure routes, to less-than-lifetime and early-life exposure, and in

combination with dioxin toxic equivalence factors. Section 5 characterizes the results

of this assessment, lists research needs, and gives specific guidance for risk

assessors.

2. SUMMARY OF STUDIES USED IN THE DOSE-RESPONSE ASSESSMENT

2.1. CANCER STUDIES IN HUMANS

EPA's cancer guidelines (U.S. EPA, 1986a, 1996a) favor basing dose-response

assessments on human studies. This requires quantitative information on both

exposure and response. This limited review focuses on the suitability of the human

studies for dose-response assessment. More detailed information on these studies and

on other studies not amenable to dose-response assessment has been compiled by

ATSDR (1993).

Bertazzi et al. (1987). This cohort study analyzed cancer mortality among

workers at a capacitor manufacturing plant in Italy. PCB mixtures with 54, then

42 percent chlorine were used through 1980. The cohort included 2100 workers (544

males and 1556 females) employed at least 1 week. At the end of follow-up in 1982,

there were 64 deaths, 26 from cancer.

In males, there was a statistically significant increase in death from

gastrointestinal tract cancer, compared with national and local rates (6 observed, 1.7

expected using national rates, SMR=346, CI=141–721; 2.2 expected using local rates,

SMR=274, CI=112–572). In females, there was a statistically significant excess risk of

death from hematologic cancer compared with local, but not national, rates (4

observed, 1.1 expected, SMR=377, CI=115–877). Analyses by exposure duration,

latency, and year of first exposure revealed no trend; however, the numbers are small.

Brown (1987). This cohort study analyzed cancer mortality among workers at

two capacitor manufacturing plants in New York and Massachusetts. At both plants the

Aroclor mixture being used changed twice, from 1254 to 1242 to 1016. The cohort

included 2588 workers (1270 males and 1318 females) employed at least 3 months in

areas of the plants considered to have potential for heavy exposure to PCBs. At the

end of follow-up in 1982, there were 295 deaths, 62 from cancer.

Compared with national rates, there was a statistically significant increase in

death from cancer of the liver, gall bladder, and biliary tract (5 observed, 1.9 expected,

SMR=263,

<0.05). Four of these five occurred among females employed at the

Massachusetts plant. Analyses by time since first employment or length of employment

revealed no trend; however, the numbers are small.

Sinks et al. (1992). This cohort study analyzed cancer mortality among workers

at a capacitor manufacturing plant in Indiana. Aroclor 1242, then 1016, had been used.

The cohort included 3588 workers (2742 white males and 846 white females) employed

at least 1 day. At the end of follow-up in 1986, there were 192 deaths, 54 from cancer.

Workers were classified into five exposure zones based on distance from the

impregnation ovens.

Compared with national rates, there was a statistically significant excess risk of

death from skin cancer (8 observed, 2.0 expected, SMR=410; CI=180–800); all were

malignant melanomas. A proportional hazards analysis revealed no pattern of

association with exposure zone; however, the numbers are small.

Other occupational studies. Other studies (NIOSH, 1977; Gustavsson et al.,

1986; Shalat et al., 1989) looked for an association between occupational PCB

exposure and cancer mortality. Because of small sample sizes, brief follow-up periods,

and confounding exposures to other potential carcinogens, these studies are

inconclusive and not amenable to dose-response analysis.

Accidental ingestion. Serious adverse health effects, including liver cancer

and skin disorders, have been observed in humans who consumed rice oil

contaminated with PCBs in the "Yusho" incident in Japan or the "Yu-Cheng" incident in

Taiwan. These effects have been attributed, at least in part, to heating of the PCBs

and rice oil, causing formation of chlorinated dibenzofurans, which have the same

dioxin-like mode of action as some PCB congeners (ATSDR, 1993; Safe, 1994).

2.2. LIFETIME CANCER STUDIES IN ANIMALS

Because of their controlled exposures and absence of confounding factors,

animal studies are often used for dose-response analysis. A new study compared

several commercial mixtures over a range of dose levels; earlier studies had focused on

mixtures with high chlorine content. This limited review focuses on the information that

would be used in a dose-response assessment. More detailed information on these

studies has been compiled by ATSDR (1993).

Kimbrough et al. (1975). Groups of 200 female Sherman rats were fed diets

with 0 or 100 ppm Aroclor 1260 for about 21 months. Six weeks later the rats were

killed and their tissues were examined. Hepatocellular carcinomas and neoplastic

nodules were significantly increased in rats fed Aroclor 1260 (see table 2–1).

Table 2–1. Liver tumor incidences in rats from lifetime exposure studies, 1975–1985

Study, sex and strain, mixture

Dose Original Reevaluation

a a,b

Kimbrough et al. (1975) Control ** 1/173 ( 1%) ** 1/187 ( 1%)

F Sherman, 1260 100 ppm 170/184 (92%) 138/189 (73%)

NCI (1978) Control ** 0/24 ( 0%) ** 0/24 ( 0%)

M Fischer, 1254 25 ppm 0/24 ( 0%) 1/24 ( 4%)

50 ppm 1/24 ( 4%) 1/24 ( 4%)

100 ppm 3/24 (12%) 3/23 (13%)

NCI (1978) Control ** 0/23 ( 0%) 0/23 ( 0%)

F Fischer, 1254 25 ppm 0/24 ( 0%) 1/24 ( 4%)

50 ppm 1/22 ( 5%) 2/24 ( 8%)

100 ppm 2/24 ( 8%) 1/24 ( 4%)

Schaeffer et al. (1984) Control ** 2/120 ( 2%) 8/120 ( 7%)

M Wistar, Clophen A 30 100 ppm 42/130 (32%) 16/128 (12%)

Schaeffer et al. (1984) Control ** 2/120 ( 2%) ** 8/120 ( 7%)

M Wistar, Clophen A 60 100 ppm 123/129 (95%) 114/125 (91%)

Norback and Weltman (1985) Control ** 0/32 ( 0%) 0/31 ( 0%)

M Sprague-Dawley, 1260 100/50/0 ppm 7/46 (15%) 5/40 (12%)

Norback and Weltman (1985) Control ** 1/49 ( 2%) ** 1/45 ( 2%)

F Sprague-Dawley, 1260 100/50/0 ppm 45/47 (96%) 41/46 (89%)

**Statistically significant (

<0.05) by Cochran-Armitage trend test (for experiments with more than one

dosed group) or Fisher exact test (for experiments with one dosed group).

Hepatocellular adenomas or carcinomas

Decreases between original and reevaluated denominators are due to lost slides; increases, to slides that

were excluded originally but could not be specifically identified for exclusion in the reevaluation.

One control group supported both experiments.

Dosing was decreased twice during the study.

Source: Adapted from Moore et al. (1994).

National Cancer Institute (NCI, 1978). Groups of 24 male or female

Fischer 344 rats were fed diets with 0, 25, 50, or 100 ppm Aroclor 1254 for

104–105 weeks (24 months). Then the rats were killed and their tissues were

examined. The combined incidence of leukemia and lymphoma in males was

significantly increased by the Cochran-Armitage trend test; however, since Fisher exact

tests were not also significant, NCI did not consider this result clearly related to

Aroclor 1254. Hepatocellular adenomas and carcinomas were increased (see table 2–1).

Morgan et al. (1981) and Ward (1985) reevaluated gastric lesions from this

study and found 6 adenocarcinomas in 144 exposed rats. This result is statistically

significant, as gastric adenocarcinomas had occurred in only 1 of 3548 control male

and female Fischer 344 rats in the NCI testing program. Intestinal metaplasia in

exposed rats differed morphologically from controls, suggesting Aroclor 1254 can act

as a tumor initiator.

Schaeffer et al. (1984). Male weanling Wistar rats were fed a standard diet for

8 weeks, then were divided into three groups. One group was fed the basic diet; for the

other groups 100 ppm Clophen A 30 or A 60 was added. Rats were killed at

801–832 days (26.3–27.3 months) and were examined for lesions in the liver and some

other tissues. For both mixtures, preneoplastic liver lesions were observed after

500 days (16.4 months) and hepatocellular carcinomas after 700 days (23 months) in

rats dying before the end of the study (see table 2–1). The investigators concluded,

"Clophen A 60 had a definite, and Clophen A 30 a weak, carcinogenic effect on rat

liver."

Norback and Weltman (1985). Groups of male or female Sprague-Dawley rats

were fed diets with 0 or 100 ppm Aroclor 1260 for 16 months; the latter dose was

reduced to 50 ppm for 8 more months. After 5 additional months on the control diet, the

rats were killed and their livers were examined. Partial hepatectomy was performed on

some rats at 1, 3, 6, 9, 12, 15, 18, and 24 months to evaluate sequential morphologic

changes. In males and females fed Aroclor 1260, liver foci appeared at 3 months, area

lesions at 6 months, neoplastic nodules at 12 months, trabecular carcinomas at

15 months, and adenocarcinomas at 24 months, demonstrating progression of liver

lesions to carcinomas. By 29 months, 91 percent of females had liver carcinomas and

95 percent had carcinomas or neoplastic nodules; incidences in males were lower, 4

and 15 percent, respectively (see table 2–1).

Vater et al. (1995) obtained individual animal results to determine whether the

partial hepatectomies, which exert a strong proliferative effect on the remaining tissue,

affected the incidence of liver tumors. They reported that the hepatectomies did not

increase the tumor incidence. Among females fed Aroclor 1260, liver tumors

developed in 4 of 7 with hepatectomies and 37 of 39 without hepatectomies; no liver

tumors developed in controls or males with hepatectomies.

Moore et al. (1994); Institute for Evaluating Health Risks (IEHR) (1991). The

preceding rat liver findings were reevaluated using criteria and nomenclature that had

changed to reflect new understanding of mechanisms of toxicity and carcinogenesis.

The reevaluation found somewhat fewer tumors than did the original investigators. The

apparent increase for Clophen A 30 (Schaeffer et al., 1984) is no longer statistically

significant. Original and revised rat liver tumor incidences are given in table 2–1.

Brunner et al. (1996). This new study compared carcinogenicity across

different Aroclors, dose levels, and sexes. Groups of 50 male or female Sprague-

Dawley rats were fed diets with 25, 50, or 100 ppm Aroclor 1260 or 1254; 50 or 100

ppm Aroclor 1242; or 50, 100, or 200 ppm Aroclor 1016. There were 100 controls of

each sex. The animals were killed at 104 weeks, after which a complete

histopathologic evaluation was performed for control and high-dose groups;

histopathologic evaluations of liver, brain, mammary gland, and male thyroid gland

were also performed for low- and mid-dose groups.

Statistically significant increased incidences of liver adenomas or carcinomas

were found in female rats for all Aroclors and in male rats for Aroclor 1260 (see

table 2–2). Several of these tumors were hepatocholangiomas, a rare bile duct tumor

seldom seen in control rats. Hepatocholangiomas occurred in three females and two

males fed 100 ppm Aroclor 1260, in two, six, and one female fed Aroclor 1254 at 25,

50, and 100 ppm, respectively, and in one and two females fed Aroclor 1242 at 50 and

100 ppm, respectively; there was a hepatocholangiocarcinoma in one female fed

50 ppm Aroclor 1242.

Table 2–2. Liver tumor incidences in rats from 1996 lifetime exposure study

Mixture

Dose Females Males

Aroclor 1260 Control ** 1/85 ( 1%) ** 7/98 ( 7%)

25 ppm 10/49 (20%) 3/50 ( 6%)

50 ppm 11/45 (24%) 6/49 (12%)

100 ppm 24/50 (48%) 10/49 (20%)

Aroclor 1254 Control ** 1/85 ( 1%) 7/98 ( 7%)

25 ppm 19/45 (42%) 4/48 ( 8%)

50 ppm 28/49 (57%) 4/49 ( 8%)

100 ppm 28/49 (57%) 6/47 (13%)

Aroclor 1242 Control ** 1/85 ( 1%) 7/98 ( 7%)

50 ppm 11/49 (24%) 1/50 ( 2%)

100 ppm 15/45 (33%) 4/46 ( 9%)

Aroclor 1016 Control ** 1/85 ( 1%) 7/98 ( 7%)

50 ppm 1/48 ( 2%) 2/48 ( 4%)

100 ppm 6/45 (13%) 2/50 ( 4%)

200 ppm 5/50 (10%) 4/49 ( 8%)

**Statistically significant (

<0.05) by Cochran-Armitage trend test.

Hepatocellular adenomas, carcinomas, cholangiomas, or cholangiocarcinomas in rats alive when the first

tumor was observed.

One control group supported all experiments.

Source: Adapted from Brunner et al. (1996), Keenan and Stickney (1996).

To investigate tumor progression after exposure stops, groups of 24 female rats

were exposed for 52 weeks, then exposure was discontinued for an additional

52 weeks before the rats were killed. For Aroclors 1254 and 1242, tumor incidences

from the stop study were approximately half those of the lifetime study; that is, nearly

proportional to exposure duration. In contrast, stop-study tumor incidences were zero

for Aroclor 1016, while for Aroclor 1260 they were generally greater than half those of

the lifetime study (see table 2–3). For 100 ppm Aroclor 1260, the stop study incidence

was greater than that of the lifetime study, 71 vs. 48 percent. (This 48 percent lifetime

study incidence was also low compared with incidences of 73, 91, and 89 percent from

the earlier studies of 100 ppm Aroclor 1260 or Clophen A 60.)

Table 2–3. Liver tumor incidences in female rats from 1996 stop study

Mixture

Dose Stop study Lifetime study

Aroclor 1260 Control ** 1/85 ( 1%) ** 1/85 ( 1%)

25 ppm 4/24 (17%) 10/49 (20%)

50 ppm 3/24 (12%) 11/45 (24%)

100 ppm 17/24 (71%) 24/50 (48%)

Aroclor 1254 Control ** 1/85 ( 1%) ** 1/85 ( 1%)

25 ppm 5/24 (21%) 19/45 (42%)

50 ppm 7/24 (29%) 28/49 (57%)

100 ppm 6/24 (25%) 28/49 (57%)

Aroclor 1242 Control ** 1/85 ( 1%) ** 1/85 ( 1%)

50 ppm 3/24 (12%) 11/49 (22%)

100 ppm 6/24 (25%) 15/45 (33%)

Aroclor 1016 Control 1/85 ( 1%) ** 1/85 ( 1%)

50 ppm 0/24 ( 0%) 1/48 ( 2%)

100 ppm 0/24 ( 0%) 6/45 (13%)

200 ppm 0/24 ( 0%) 5/50 (10%)

**Statistically significant (

<0.05) by Cochran-Armitage trend test.

Hepatocellular adenomas, carcinomas, or cholangiomas in female rats dosed for 52 weeks and killed at

104 weeks.

Hepatocellular adenomas, carcinomas, cholangiomas, or cholangiocarcinomas in female rats dosed for

104 weeks and killed at 104 weeks (comparison from table 2–2).

One control group supported all experiments.

Source: Adapted from Brunner et al. (1996).

Thyroid gland follicular cell adenomas or carcinomas were increased in males

for all Aroclors (see table 2–4); significant dose trends were noted for Aroclors 1254

and 1242. The increases did not continue proportionately above the lowest dose. No

trends were apparent in females.

Table 2–4. Thyroid gland tumor incidences in male rats from 1996 lifetime exposure study

Mixture

Dose Males

Aroclor 1260 Control 2/100 ( 2%)

25 ppm 7/50 (14%)

50 ppm 5/50 (10%)

100 ppm 4/50 ( 8%)

Aroclor 1254 Control ** 2/100 ( 2%)

25 ppm 7/50 (14%)

50 ppm 7/50 (14%)

100 ppm 6/50 (12%)

Aroclor 1242 Control ** 2/100 ( 2%)

50 ppm 7/50 (14%)

100 ppm 6/50 (12%)

Aroclor 1016 Control 2/100 ( 2%)

50 ppm 4/50 ( 8%)

100 ppm 3/50 ( 6%)

200 ppm 1/50 ( 2%)

**Statistically significant (

<0.05) by Cochran-Armitage trend test.

Follicular cell adenomas or carcinomas in male rats dosed for 104 weeks.

One control group supported all experiments.

Source: Adapted from Brunner et al. (1996).

In female rats, the incidence of mammary tumors was decreased with lifetime

exposure to Aroclor 1254 and, to a lesser extent, to 1260 or 1242; this result was not

observed for Aroclor 1016. Decreases did not occur for any Aroclor in the stop study.

The first mammary tumor was observed at a later age in the dosed groups.

Studies of structurally related agents. Studies of 2,3,7,8-

tetrachlorodibenzo-p-dioxin and a polybrominated biphenyl (PBB) mixture are

summarized here because the pattern of tumors found by Brunner et al. (1996) mimics

the tumors induced in rats by these structurally related agents.

The National Toxicology Program (NTP, 1982) exposed groups of 50 male or

female Osborne-Mendel rats by gavage to 0, 1.4, 7.1, or 71 ng/kg-d 2,3,7,8-

tetrachlorodibenzo-p-dioxin for 2 years. Similar to the Brunner et al. (1996) study, liver

tumors were increased in female rats and thyroid gland follicular cell tumors were

increased in male rats. Mammary tumors were not, however, decreased in dosed

female rats.

NTP (1983) exposed groups of 51 male or female Fischer 344/N rats by gavage

to 0, 0.1, 0.3, 1, 3, or 10 mg/kg-d of a PBB mixture ("Firemaster FF–1") for 6 months,

then exposure was discontinued for 23 months before the animals were killed.

Statistically significant increased incidences of liver tumors were found in male and

female rats. Dose-related increased incidences of cholangiocarcinomas were found in

male and female rats. The Firemaster FF–1 mixture comprised an anticaking agent

blended with a PBB mixture containing 56 percent 2,4,5,2',4',5'-hexabromobiphenyl,

27 percent 2,3,4,5,2',4',5'-heptabromobiphenyl, and other unspecified penta-, hexa-,

and heptabromobiphenyls. The analogous PCB congeners are noted for their high

toxicity and abundance in environmental samples (McFarland and Clarke, 1989);

2,4,5,2',4',5'-hexachlorobiphenyl is highly persistent in the body (Matthews and

Anderson, 1975) and comprises 21.5 and 12.0 percent, respectively, of PCB residues

in human fat and milk (McFarland and Clarke, 1989).

2.3. PARTIAL LIFETIME STUDIES IN ANIMALS

Although lifetime studies are preferred for dose-response modeling, partial

lifetime studies often use experimental designs addressing specific issues in the

application of a dose-response assessment. Partial lifetime studies for PCBs have

compared different commercial mixtures and the relative sensitivity of the sexes. Some

studies examined early-life exposure, which is not covered by most lifetime cancer

studies, where exposure starts at age 2–3 months, when the animals are mature. This

limited review focuses on the information that pertains to issues in the dose-response

assessment. More detailed information on these studies has been compiled by ATSDR

(1993).

Kimbrough et al. (1972). Groups of 10 male or female Sherman rats were fed

diets with 0, 20, 100, 500, or 1000 ppm Aroclor 1254 or 1260, beginning at 3–4 weeks

of age and continuing for 8 months. Incidences of adenofibrosis reached 2/10 in males

and 4/7 in females fed 1000 ppm Aroclor 1260; in contrast, for 100 and 500 ppm

Aroclor 1254, incidences were 1/10 and 10/10 in males and 7/10 and 9/9 in females.

There was no adenofibrosis in 10 controls of each sex. With regard to differences

between sexes, the investigators concluded Aroclor 1260 is more toxic to female rats

than males, but such a difference could not be established for Aroclor 1254. With

regard to differences between mixtures, the investigators concluded the effect on the

liver "is more pronounced with Aroclor 1254 when all morphologic changes of

equivalent dietary levels of Aroclor 1254 and 1260 are compared."

Although adenofibromas are not carcinomas, these lesions, particularly in less-

than-lifetime studies, are sometimes regarded as indicating a potential for tumor

formation over a longer duration. For example, in a subsequent study, most female rats

of this strain fed 100 ppm Aroclor 1260 developed hepatocellular carcinomas or

neoplastic nodules after 23 months (Kimbrough et al., 1975).

Kimbrough and Linder (1974). Groups of 50 male BALB/cJ mice were fed

diets with 300 ppm Aroclor 1254 for 11 months, or for 6 months followed by 5 months

without exposure. Hepatomas were found in 9 of 22 surviving mice exposed for

11 months, in 1 of 24 mice exposed for 6 months, and in none of 58 controls.

Adenofibrosis was observed in all mice exposed for 11 months, but in none of the

others.

Kimura and Baba (1973). Groups of 10 male or female Donryu rats were fed

diets that increased from 38 to 462 ppm (time-weighted average, 330 ppm)

Kanechlor 400, beginning at 10 weeks of age and continuing for different durations of

up to 400 days (13 months). Multiple adenomatous liver nodules were found in the six

females exposed for the longest durations. No nodules were found in males or in five

controls of each sex.

Ito et al. (1973). Groups of 12 male dd mice were fed diets with 100, 250, or

500 ppm Kanechlor 300, 400, or 500, beginning at 8 weeks of age and continuing for

32 weeks (7.5 months). Among mice fed 500 ppm Kanechlor 500, five had

hepatocellular carcinomas and seven had nodular hyperplasia. No other groups,

including six controls, showed these effects.

Ito et al. (1974). Male Wistar rats were fed diets with 0, 100, 500, or 1000 ppm

Kanechlor 300, 400, or 500, beginning at 8 weeks of age and continuing for

28–52 weeks (6.5–12 months). Nodular hyperplasia was seen with all three mixtures,

highest for Kanechlor 500 and lowest for Kanechlor 300, but not in controls.

Histologically, the nodular hyperplasia was similar to that induced by other chemical

carcinogens, suggesting the nodular hyperplasia is preneoplastic. The investigators

concluded, "Hepatocellular carcinomas could be induced by administration of

Kanechlor-500, -400, or -300 for a longer period."

Rao and Banerji (1988). Groups of 32 male Wistar rats were fed diets with 0,

50, or 100 ppm Aroclor 1260, beginning at 5 weeks of age and continuing for 120 days

(4 months). Neoplastic nodules with adenofibrosis were found in 24 of 32 rats fed 50

ppm Aroclor 1260 and in 16 of 32 rats fed 100 ppm. None of 32 controls showed these

changes. The investigators concluded Aroclor 1260 induces liver tumors when fed to

young rats for a short time.

2.4. TUMOR INITIATING AND PROMOTING ACTIVITY

Studies of tumor initiating and promoting activity are available for a few

commercial mixtures and congeners. The congener studies are beginning to identify a

subset of mixture components that may be significant contributors to cancer induction.

As some of these congeners are present in environmental mixtures, these studies

provide information about the potential for environmental mixtures to cause cancer.

This limited review focuses on identifying congeners with tumor promoting activity to

help risk assessors know what to look for in a site-specific congener analysis. More

detailed information on these and other studies has been compiled by Silberhorn et al.

(1990).

Several commercial PCB mixtures and congeners show tumor promoting activity

(Silberhorn et al., 1990). Aroclor 1254 and Kanechlors 400 and 500 promote liver

tumors in initiation-promotion studies; Aroclor 1254 also promotes lung tumors

(Anderson et al., 1983, 1994; Beebe et al., 1992, 1993). Aroclor 1254, Clophens A 30

and A 50, four tetrachlorobiphenyls, three pentachlorobiphenyls, and one

hexachlorobiphenyl showed promoting activity in studies to identify alterations in

adenosine triphosphatase (ATPase), gamma-glutamyl transpeptidase (GGT), or

placental glutathione S-transferase (PGST) activity, markers of tumor promoting activity

in the liver. One study found the interaction of 2,5,2',5'- and 3,4,3',4'-

tetrachlorobiphenyl to produce more alterations than either alone (Sargent et al., 1991).

One monochlorobiphenyl and one dichlorobiphenyl showed no promoting activity. Lists

of mixtures and congeners tested for promoting activity (with either positive or negative

results) appear in table 2–5; references can be found in Silberhorn et al. (1990) and

later references cited in the table.

Table 2–5. Mixtures and congeners tested for tumor promoting activity

Mixture

Tumors Mixture or congener Altered foci

Aroclor 1254 Liver, lung Aroclor 1254 GGT+

Kanechlor 400 Liver Clophen A 30 Marker not reported

Kanechlor 500 Liver Clophen A 50 ATPase-, GGT+

4–MCB Negative

4,4'–DiCB Negative

2,4,2',4'–TeCB GGT+

2,4,2',5'–TeCB ATPase–, GGT+

2,5,2',5'–TeCB ATPase–, PGST+

3,4,3',4'–TeCB ATPase–, GGT+, PGST+

2,3,4,3',4'–PeCB GGT+, PGST+

2,4,5,3',4'–PeCB ATPase–, GGT+

3,4,5,3',4'–PeCB GGT+, PGST+

2,4,5,2',4',5'–HxCB ATPase–, GGT+, PGST+

Compiled from many studies; not all mixtures or congeners were tested in all systems.

Sources: Adapted from Silberhorn et al. (1990), Buchmann et al. (1991), Laib et al. (1991), Sargent et al.

(1991), Beebe et al. (1992, 1993), Hemming et al. (1993), Anderson et al. (1994).

Although PCBs are not generally described as tumor initiators, in some studies a

small number of ATPase-deficient or GGT-positive foci were initiated by treatment with

Clophen A 50 alone (Silberhorn et al., 1990). Weak initiating activity was found with

2,4,2',5'-tetrachlorobiphenyl, which induced ATPase-deficient, but not GGT-positive,

foci (Rose et al., 1985; Laib et al., 1991). Initiation potential had been suggested by

the different intestinal metaplasia morphology induced by Aroclor 1254 (Morgan et al.,

1981; Ward, 1985). Many other investigators, however, report negative results for

tumor initiation by PCB mixtures or congeners (Silberhorn et al., 1990).

The significance of the promotion studies is apparent, as all six ortho-substituted

congeners producing altered foci are abundant in commercial mixtures (Schulz et al.,

1989) and have been found in environmental samples (Lake et al., 1995; McFarland

and Clarke, 1989), though the tetrachlorobiphenyls are not particularly persistent in the

environment. Known for its bioaccumulation potential and abundance in environmental

samples, 2,4,5,2',4',5'-hexachlorobiphenyl has been found to comprise 21.5 and 12.0

percent, respectively, of PCB residues in human fat and milk; 2,4,5,3',4'-

pentachlorobiphenyl constitutes 5.4 and 6.5 percent, respectively, of these residues

(McFarland and Clarke, 1989). The coplanar congeners 3,4,3',4'-tetrachlorobiphenyl

and 3,4,5,3',4'-pentachlorobiphenyl have lower abundance in commercial mixtures

(Kannan et al., 1988) but have been found in a variety of organisms, including humans

(Safe, 1994).

2.5. ABSORPTION AND RETENTION

Cancer studies of lifetime and partial lifetime PCB exposure have been by

ingestion only. Pharmacokinetic studies provide information about the potential for

absorption and a risk of cancer by other exposure routes. Other studies have quantified

the retention and persistence of PCBs in the body. This limited review focuses on the

information that pertains to applying the dose-response assessment to dermal and

inhalation exposure. More detailed information on these studies and on ingestion

studies has been compiled by ATSDR (1993).

Humans absorb PCBs from ingestion, inhalation, and dermal exposure (ATSDR,

1993). Once absorbed, PCBs enter the circulation and are transported throughout the

body. Initial distribution is to liver and muscle, which are highly perfused;

subsequently, PCBs, being highly lipophilic, accumulate in fat and skin (Matthews and

Anderson, 1975).

Inhalation can be a principal absorption route for occupational PCB exposure

(Wolff, 1985). In animals, an inhaled PCB aerosol was rapidly absorbed, although

rates were not estimated (ATSDR, 1993).

PCBs can cross human skin and increase the body burden. Dermal exposure

can contribute significantly to body burdens of workers (Wolff, 1985) and can be a

major route of environmental exposure (ATSDR, 1993). In vivo dermal absorption by

rhesus monkeys exposed for 24 hours to soil containing 44 ppm Aroclor 1242 or

23 ppm Aroclor 1254 was 14 percent in each case (Wester et al., 1993). Earlier

studies found similar absorption rates for PCBs in mineral oil, trichlorobenzene, and

acetone (Wester et al., 1990). Subsequent washing did not remove all PCBs,

especially if time had elapsed after exposure (Wester et al., 1983). In vitro human skin

accumulation of Aroclor 1254 from water was 12 percent after a half hour (Wester

et al., 1987) and 44 percent after 24 hours (Wester et al., 1990), suggesting absorption

is rapid initially and continues at a slower rate with further contact.

PCBs are eliminated through metabolism, which occurs primarily in the liver

(Matthews and Anderson, 1975). Metabolism rates are generally lower with high

chlorine content, but chlorine position is also important (Hutzinger et al., 1974;

Matthews and Anderson, 1975). Absence of chlorine at two adjacent positions

facilitates metabolism (Matthews and Anderson, 1975). Metabolism and elimination

can be quite slow; for example, the biological half-life of 2,4,5,2',4',5'-

hexachlorobiphenyl exceeds the lifespan of rats (Matthews and Anderson, 1975).

In addition to variability by congener, there is human variability in PCB

metabolism and elimination. People with decreased liver function, including inefficient

glucuronidative mechanisms in infants, can have less capacity to metabolize and

eliminate PCBs (Calabrese and Sorenson, 1977). Additionally, approximately five

percent of nursing infants receive a steroid in human milk that inhibits the activity of

The workers came from the plant studied by Sinks et al. (1992), where Aroclor 1242, then 1016, had been used. The

quantitation of PCBs as Aroclors 1254 and 1242 illustrates both (1) selective retention of congeners with high chlorine

content and (2) the imprecision of characterizing altered PCB mixtures as if they were Aroclors.

Serum concentration was modeled as an exponentially decreasing function of time:

exp(–

), where

concentration at time

is initial concentration, and

is the rate parameter, estimated by linear regression of ln(

) on

0 0

glucuronyl transferase, further reducing PCB metabolism and elimination (Calabrese

and Sorenson, 1977).

Persistent congeners can retain biological activity long after exposure stops;

residual liver enzyme induction was observed in mice 42 weeks after a single dose of

Aroclor 1254 (Anderson et al., 1991a). The majority of the retained mixture comprised

2,4,5,3'4'- and 2,3,4,3',4'-pentachlorobiphenyl and 2,4,5,2',4',5'- and 2,3,4,2',4',5'-

hexachlorobiphenyl (see tables 2–5, 3–3, and 3–4).

Analysis of 1977 and 1985 serum levels in 58 Indiana workers exposed to PCBs

yielded median half-lives of 2.6 years for Aroclor 1242 and 4.8 years for Aroclor 1254

(Phillips et al., 1989). Among workers with lowest concentrations (0–30 ppb), median

half-lives were higher, 3.1 years for Aroclor 1242 and 6.5 years for Aroclor 1254. In

another study in the same Indiana city, from 1977 to 1984 serum levels in five workers

exposed to PCBs decreased 89–94 percent (median, 92 percent) for Aroclor 1242 and

14–53 percent (median, 16 percent) for Aroclor 1260; among six others without current

occupational exposure, decreases were 23–71 percent (median, 39 percent) for total

PCBs (Steele et al., 1986). Analysis of serum levels in the exposed workers yields

half-lives of 2 years for Aroclor 1242 and 16 years for Aroclor 1260; in those without

current occupational exposure, a half-life of 8 years for total serum PCBs.

A study of people exposed through eating contaminated fish suggests that these

mixtures can be more persistent. From 1977 to 1985 mean serum levels (quantified

using Aroclor 1260 as a reference standard) from 111 Great Lakes fish eaters

decreased only slightly, from 20.5 to 19.0 ppb (Hovinga et al., 1992).

To illustrate, consider a mixture of two components in equal parts: one component has a half-life of 1 year; the other,

100 years. If the mixture concentration is sampled after 10 years, the half-life of the total mixture will appear to be

approximately 10 years: virtually all the first component will be gone, virtually none of the second, so about half the

original mixture will remain. This half-life, however, overestimates the slow rate of decrease in the more persistent

mixture fraction that remains.

It is important to recognize that ascribing a half-life to a mixture is problematic if

half-lives of its components differ widely; more specifically, half-life estimates for a

mixture can underestimate its long-term persistence.

2.6. METABOLISM AND MODE OF ACTION IN THE LIVER

Mechanistic information provides insight and understanding of the biological

activity of PCBs and their metabolites. The following discussion was contributed by

peer reviewers Drs. Larry Robertson and Lucy Anderson.

Although the rate of metabolism is slow (Mills et al., 1985), PCBs may be

converted by hepatic enzymes to hydroxylated metabolites. The relative rates of

conversion are dependent on the number and placement of the chlorine atoms present.

PCBs with fewer chlorines and with adjacent, unsubstituted carbon atoms are more

readily susceptible to metabolic attack. Cytochrome P–450 isozymes may catalyze

these hydroxylation reactions via an electrophilic arene oxide intermediate or via direct

insertion mechanisms. Evidence for the intermediacy of arene oxides during PCB

metabolism is found in the identification of (1) NIH-shift products, (2) dihydrodiol

metabolites, (3) mercapturic acid products, and (4) sulfone metabolites (Sipes and

Schnellman, 1987).

PCB metabolites with multiple hydroxyl groups also have been identified in

animals and in microsomal incubations (McLean et al., 1996a). Dihydroxy metabolites

may be oxidized in vitro to o- or p-quinones by peroxidases. In vitro studies have

demonstrated that adducts of PCBs and nucleotides (dGp and dAp) or exogenous DNA

may be formed during the hydroxylation step (from electrophilic arene oxides) and

during the peroxidase-catalyzed oxidation of PCB catechol and hydroquinone

metabolites to the respective o- and p-quinones (McLean et al., 1996b; Oakley et al.,

1996). Hydroxylated PCB metabolites may have estrogenic activity (Gierthy et al.,

1995).

Higher halogenated PCBs may be efficacious inducers of xenobiotic-

metabolizing enzymes, although they are poor substrates. Several PCBs, possessing

no or one ortho chlorine, bind the aryl hydrocarbon receptor with avidity (Bandiera

et al., 1982) and induce cytochrome P–450 1A. Several di-ortho substituted PCBs

induce cytochrome P–450s as does phenobarbital, while other congeneric PCBs may

induce cytochrome P–450s from both subfamilies. Many of these PCBs may also

induce epoxide hydrolase, glutathione transferases, and glucuronosyl transferases.

Induction of xenobiotic metabolites may be accompanied by an increase in hepatic cell

size and number and a proliferation of the endoplasmic reticulum. The persistent

induction of hepatic cytochrome P–450s, in the absence of an oxidizable xenobiotic

substrate, may provide suitable conditions for generation of reactive oxygen species.

Several PCBs tested as promoters in rat two-stage hepatocarcinogenesis were

efficacious when they were administered at doses that caused liver hypertrophy and

the induction of cytochrome P–450s (Silberhorn et al., 1990). Promoter activity has

been observed among groups of PCB congeners that have been characterized as

having widely different kinds of biological activity, including congeners that are aryl

hydrocarbon agonists, congeners that induce cytochrome P–450 1A and 2B isozymes,

and congeners that have a pattern of enzyme induction similar to that of phenobarbital.

This may indicate multiple mechanisms of action for promotion (Buchmann et al.,

1991). Congeneric PCBs may interfere with gap-junctional intercellular communication

via structure-specific mechanisms. Mono- and di-ortho chlorine substituted PCBs were

more active (Swierenga et al., 1990).

2.7. MODE OF ACTION IN THE THYROID

Recent mechanistic insights into thyroid carcinogenesis provide a rationale for

choosing a dose-response approach for thyroid tumors. The following discussion was

contributed by EPA consensus reviewer Dr. Richard Hill.

Thyroid tumors are noted for PCBs (Brunner et al., 1996) and for structurally

related 2,3,7,8-tetrachlorodibenzo-p-dioxin (NTP, 1982). These compounds are

accompanied by a lack of mutagenic activity in many different test systems. Liver

microsomal enzyme inducers of both the AHH and PB types, which include the PCBs,

commonly increase the metabolism and excretion of thyroid hormone (McClain, 1989).

Depending upon the compound, there may be increased clearance of thyroid hormone

from the blood, accentuated binding of the hormone in the liver, increased

glucuronidation of the hormone following induction of UDP-glucuronyl transferase

(Barter and Klaassen, 1992), increased bile flow and increased excretion of hormone in

the bile. Effects are usually more pronounced for thyroxine (T4) than triiodothyronine

(T3).

PCBs have effects on thyroid hormone status independent of their influence on

thyroid hormone metabolism and excretion. They cause damage to follicular cells

(Kasza et al., 1978; Byrne et al., 1987) and bind to and possibly displace thyroid

hormone from plasma protein carriers (Rickenbacher et al., 1986). Both of these may

contribute to a reduction in effective levels of circulating thyroid hormone.

Decreases in circulating thyroid hormone stimulate the pituitary by negative

feedback to increase the output of thyroid stimulating hormone (TSH). TSH is a trophic

hormone for the thyroid, resulting in the increased synthesis of thyroid hormone. When

thyroid hormone needs cannot be met by existing follicular cells, cells undergo

hypertrophy and diffuse hyperplasia. With continuing disruption in thyroid-pituitary

status, focal hyperplasia and then benign and malignant neoplasms develop (Hill et al.,

1987).

It is not totally clear whether hormonal derangement noted in rodents is a factor

in the development of thyroid tumors in humans. Some studies of persons with iodide

deficiency or inborn deficiency in the synthesis of thyroid hormone support the

contention, while others do not. At this time there is not enough information to dismiss

the animal model as not being relevant to human thyroid carcinogenesis. Even if

humans are susceptible to cancer from thyroid-pituitary disruption, existing information

indicates that humans are less sensitive than rodents (Hill et al., 1987).

Thyroid cancer risks in rodents exist under conditions of disruption in thyroid-

pituitary status. When, however, circulating levels of thyroid hormone and TSH pertain,

risks would be expected to be minimal. Such findings are best expressed by nonlinear

dose-response relationships. In assessing the risks from thyroid tumors, one would

want dose-response and time-action data from repeat dose studies on such things as

thyroid weight and morphology, UDP-glucuronyl transferase activity, and thyroid

hormone and TSH levels. Points of departure for evaluation of risks could be

determined from doses not associated with perturbations in thyroid status. Margin of

exposure—the ratio of the point of departure to expected human exposure

levels—could be used to express the nonlinear risks.

3. DOSE-RESPONSE ASSESSMENT

3.1. APPROACHES TO DOSE-RESPONSE ASSESSMENT

Dose-response assessment begins with consideration of developing a

biologically based model, that is, a model whose mathematical structure reflects the

ascertained mode of action and whose parameters are measured in experimental

studies. Biologically based models have been developed for 2,4,2',5'- and 3,4,3',4'-

tetrachlorobiphenyl; few congeners or mixtures, however, have been tested to measure

the rate parameters that would be used in a biologically based model. Further, PCBs

can cause cancer through multiple modes of action (Safe, 1990, 1994), indicating a

need for multiple models. Consequently, the information available at this time is more

suited to empirical modeling, where a flexible default model—allowing either linearity or

nonlinearity—is fitted to describe tumor incidence as a function of dose in the

experimental range.

Extrapolation to lower doses considers both linear and nonlinear approaches,

with a linear default if there is not sufficient information to support a sublinear model

(U.S. EPA, 1986a, 1996a). This policy rests, in part, on some general considerations.

Low-dose-linear models are appropriate for extrapolation to lower doses when a

carcinogen acts in concert with other exposures and processes that cause a

background incidence of cancer (Crump et al., 1976; Lutz, 1990). Further, even when

the mode of action indicates a nonlinear dose-response curve in homogeneous animal

populations, the presence of genetic and lifestyle factors in a heterogeneous human

population tends to make the dose-response curve more linear (Lutz, 1990). This is

because genetic and lifestyle factors contribute to a wider spread of human sensitivity,

which extends and straightens the dose-response curve over a wider range. Although

these considerations provide a reasonable argument for a model that is linear at low

doses, the relation of the low-dose slope to one from the experimental range is

uncertain; this uncertainty increases with the distance from the experimental range.

PCBs give generally negative results in tests of genetic activity (ATSDR, 1993),

implying that PCBs induce tumors primarily through modes of action that do not involve

gene mutation. This raises the possibility of a nonlinear dose-response curve. There

is, however, no dose-response information on either tumors or tumor precursors to

describe the dose range where the curve would be sublinear. At the low end of the

experimental range (25–100 ppm), dose-response curves are not sublinear, as tumor

incidence declines less than proportionately with dose for Aroclors 1260, 1254, and

1242 (Brunner et al., 1996). At much lower doses, some PCB congeners add to the

considerable background of human exposure to dioxin-like compounds and augment

processes associated with dioxin toxicity, providing a linear component to the dose-

response curve. Between these supralinear and linear dose ranges there is no dose-

response information; consequently, the information available at this time is more

suited to linear extrapolation.

Environmental PCBs occur as mixtures, prompting consideration of which agents

provide the most appropriate basis for an assessment. EPA's mixture guidelines (U.S.

EPA, 1986b) favor basing assessments on the effects of the mixture of interest; the

second choice is to use a sufficiently similar mixture; next, to assess the components of

the mixture. The guidelines further advise,

Attention should also be given to the persistence of the mixture in the

environment as well as to the variability of the mixture composition over

time or from different sources of emissions. If the components of the

mixture are known to partition into different environmental compartments

or to degrade or transform at different rates in the environment, then

those factors must also be taken into account, or the confidence in and

applicability of the risk assessment is diminished.

There are no cancer studies of PCB mixtures found in the environment. Studies

are available for some commercial mixtures, though their similarity to an environmental

mixture can be a matter of considerable uncertainty as mixtures are partitioned,

transformed, and bioaccumulated in the environment. Assessing mixture components

is not now a viable alternative, because only a few congeners have been tested, none

in long-term carcinogenesis studies. Thus assessments of environmental mixtures

must use information on commercial mixtures. Partitioning, transformation, and

bioaccumulation in the environment, however, must also be taken into account.

Risk estimates can be derived from either human or animal studies; each has

strengths and limitations. Estimates derived from human studies reflect an observed

association between human exposure and cancer; however, it is difficult to reconstruct

reliable estimates of past exposure and separate the effect of confounding exposures

to other carcinogens. Estimates derived from animal studies benefit from controlled

exposures and absence of confounding factors; however, there is uncertainty in

extrapolating dose and response rates across species. EPA's cancer guidelines (U.S.

EPA, 1986a, 1996a) favor basing dose-response assessments on human studies. In

the absence of adequate human information, assessments use animal species

responding most like humans. If this cannot be determined, assessments emphasize

long-term animal studies showing the greatest sensitivity, with due regard to biological

relevance, exposure route, and statistical considerations; this default is considered to

be conservative, tending toward public health protection.

For PCBs, the human studies involve relatively few cancer cases and lack

contemporaneous exposure estimates. Some studies report air concentrations, but

because skin contact is a major route of occupational exposure, air concentrations

would be a poor measure of exposure (Bertazzi et al., 1987; Brown, 1987). Some

studies report blood levels, but for relatively few workers at the end of exposure

(Bertazzi et al., 1987; Brown, 1987; Taylor, 1988; Sinks et al., 1992). Reconstruction of

past exposure is problematic because different mixtures had been in use over the

years, the distribution of exposure and absorption by route and congener is unknown,

and congener persistence in the body varies greatly from congener to congener

(Brown, 1994) and person to person (Steele et al., 1986). Similarly, adjustment for

confounding exposures to other potential carcinogens, many of them unidentified, is

also problematic. Because of these limitations in quantitative information, the human

studies are not well suited to dose-response assessment. Because of their controlled

exposures, absence of confounding factors, and ability to provide comparable

information on a range of different mixtures, the animal studies will be used for dose-

response modeling.

A biologically based model for two congeners is discussed in section 3.2.

Empirical models are developed and discussed in section 3.3. Analyses of congener

toxicity are discussed in section 3.4.

3.2. BIOLOGICALLY BASED MODELING OF TUMOR PROMOTION

Using a two-stage carcinogenesis model, Luebeck et al. (1991) modeled tumor

promoting activity of 2,4,2',5'- and 3,4,3',4'-tetrachlorobiphenyl, based on the study of

Buchmann et al. (1991). Female Wistar rats were initiated with 10 mg/kg-d

diethylnitrosamine for 10 days, followed by eight weekly injections of 10 or 150 mol/kg

of various compounds. The rats were killed 1 or 9 weeks later, and preneoplastic

activity was characterized by changes in ATPase and GGT activity.

Because results are available for two times after dosing stopped, modeling can

assess persistence of promoting activity. There was little or no promoting activity by

Formerly, EPA guidelines (U.S. EPA, 1986a) called for using the linearized multistage procedure, which fits models

where risk is a function of dose

Risk(

) = 1 – exp(–

–

– . . . –

);

=1,...,

The linearized multistage procedure fits several such models, through degree

=6. The peer review panel objected to

the polynomial degree exceeding the number of dosed groups (U.S. EPA, 1996b). Consequently, this assessment uses

linear-quadratic models (that is,

=2) unless there is only one dosed group, in which case it uses a linear model (

=1).

Equivalent human dose = animal dose × (animal weight / 70 kg human weight) .

7 1/4

The tables compiled by Brunner et al. (1996) combined hepatocholangiomas with hepatocellular adenomas and

carcinomas, a combination not recommended by McConnell et al. (1986). Individual animal results needed to remove

the hepatocholangiomas from the incidences used for modeling were not provided to EPA. The effect, however, is

expected to be negligible, as few rats had a hepatocholangioma.

2,4,2',5'-tetrachlorobiphenyl after dosing stopped, but 3,4,3',4'-tetrachlorobiphenyl

continued to promote vigorously (Luebeck et al., 1991).

Modeling can also estimate the probability of altered foci becoming extinct; that

is, disappearing after dosing stops. After dosing stops, the probability of extinction is

high, as most altered foci do not develop into observable tumors. Those that become

large, however, tend to persist, as the probability of extinction decreases as size

increases (Luebeck et al., 1991).

3.3. EMPIRICAL MODELING OF TUMOR INCIDENCE

The new feeding study by Brunner et al. (1996), with parallel experiments for

four commercial mixtures in both sexes of rats, provides the most comprehensive

information for empirical modeling. Each experiment tested several dose levels,

providing information about the shape of the dose-response curve in the experimental

range. For each mixture and sex, a linear-quadratic multistage model (Howe et al.,

1986) was fitted to experimental results. Dose was expressed as a lifetime daily

average (U.S. EPA, 1986a, 1992a), calculated from weekly body weight measurements

and food consumption estimates (Keenan and Stickney, 1996). Doses were scaled to

humans using a factor based on the 3/4 power of relative body weight (U.S. EPA,

1992b). Response was taken as the incidence of hepatocellular adenomas or

carcinomas ; combining adenomas and carcinomas reflects guidance of the National

The slope is the change in response divided by the change in dose. Relative to the origin (an increased response of 0

at a dose of 0), the change in response is 0.10 at the dose ED10; thus the slope is 0.10/ED10.

Toxicology Program (McConnell et al., 1986) and the observed progression of

hepatocellular adenomas to carcinomas (Norback and Weltman, 1985).

Cancer potency is described by an ED10 (estimated dose associated with

10 percent increased incidence) and its lower bound, LED10. ED10s have been used

both for potency ranking and as a starting point for low-dose extrapolation (Cogliano,

1986; U.S. EPA, 1988b, 1994a; National Research Council, 1993). These measures

are expressed as equivalent human doses.

For extrapolation to lower doses, an ED10 can be converted to a slope by

computing 0.10/ED10. (Note that slopes are inversely proportional to ED10s; high

potency is indicated by high slopes, but low ED10s.) Similarly, an upper-bound slope

can be obtained by computing 0.10/LED10. Formerly, upper-bound slopes were

calculated by the linearized multistage procedure (U.S. EPA, 1980, 1986a); these are

reported in the appendix as "

*"s. The LED10 method and the linearized multistage

procedure give similar upper-bound slopes; for example, for female rats fed

Aroclor 1254, the LED10 method and the linearized multistage procedure give upper-

bound slopes of 1.5 and 1.6 per mg/kg-d, respectively. Potency and slope estimates

are compiled in table 3–1; details supporting the calculations appear in appendix tables

A–1 through A–8.

Equivalent human dose (mg/kg-d) = (ppm in diet) × 0.05 × (animal weight / 70 kg human weight) .

10 1/4

In the studies by Kimbrough et al. (1975) and Norback and Weltman (1985) initial dose levels were later decreased or

discontinued. It is likely, however, that tumor development had already begun and internal exposure remained high with

release of PCBs stored in fat (Vater et al., 1995). Thus, for these studies, initial dose levels were used without averaging

over the study duration; this reduces these potency estimates by up to one-third compared to the default of averaging

dose over the study duration.

Table 3–1. Human potency and slope estimates derived from rat liver tumors

Central Upper-bound See

Study, sex and strain, mixture

ED10 LED10 slope slope table

abcd

Brunner, F Sprague-Dawley, 1260 0.24 0.19 0.4 0.5 A–1

Brunner, F Sprague-Dawley, 1254 0.086 0.067 1.2 1.5 A–2

Brunner, F Sprague-Dawley, 1242 0.38 0.27 0.3 0.4 A–3

Brunner, F Sprague-Dawley, 1016 2.4 1.4 0.04 0.07 A–4

Brunner, M Sprague-Dawley, 1260 1.0 0.55 0.1 0.2 A–5

Brunner, M Sprague-Dawley, 1254 1.7 0.87 0.06 0.1 A–6

Brunner, M Sprague-Dawley, 1242 2.9 1.2 0.03 0.08 A–7

Brunner, M Sprague-Dawley, 1016 5.9 2.5 0.02 0.04 A–8

Kimbrough, F Sherman, 1260 0.10 0.091 1.0 1.1 A–9

NCI, M Fischer, 1254 1.0 0.55 0.1 0.2 A–10

NCI, F Fischer, 1254 1.2 0.61 0.08 0.2 A–11

Schaeffer, M Wistar, A 30 2.1 1.0 0.05 0.1 A–12

Schaeffer, M Wistar, A 60 0.058 0.047 1.7 2.1 A–13

Norback, M Sprague-Dawley, 1260 1.0 0.53 0.1 0.2 A–14

Norback, F Sprague-Dawley, 1260 0.062 0.046 1.6 2.2 A–15

Estimated dose associated with 10% increased incidence, in mg/kg-d.

95% lower bound on ED10, in mg/kg-d.

Per mg/kg-d, computed as 0.10/ED10.

Per mg/kg-d, computed as 0.10/LED10.

No significant increase; quantities indicate sensitivity of study.

In conjunction with the Brunner et al. (1996) study, the earlier studies provide

useful information on the potential for lot-to-lot and strain-to-strain differences.

Because these studies did not report body weight and food consumption, administered

doses were converted from ppm in the diet to mg/kg-d using default factors based on

rats weighing 350 grams and consuming food equal to 5 percent of body weight daily

(U.S. EPA, 1980). Potency and slope estimates from the earlier studies are included

in table 3–1; details, in tables A–9 through A–15.

The range of potency values in table 3–1 is summarized in table 3–2. It is based

primarily on the range for Aroclors 1260, 1254, 1242, and 1016 in female Sprague-

Dawley rats (Brunner et al., 1996), but considers the other studies, too. For example,

the two studies of female Sprague-Dawley rats fed Aroclor 1260 (Brunner et al., 1996;

Norback and Weltman, 1985) could reflect, in part, lot-to-lot differences that are

pertinent to mixtures altered in the environment; thus the range of upper-bound slopes

includes those from the Brunner et al. (1996) study (0.07–1.5 per mg/kg-d) and the

Norback and Weltman (1985) study (2.2 per mg/kg-d). The earlier gastric tumors and

leukemias and lymphomas in male rats fed Aroclor 1254 (NCI, 1978) are not included

in this range, because they were not confirmed by the Brunner study; they would

contribute little to the overall estimates since incidences are several-fold less than

those of the liver tumors.

Table 3–2. Range of human potency and slope estimates

Central Upper-bound See

ED10

LED10 slope slope table

abcd

Highest observed potency 0.086 0.046 1.2 2.2 A–2,15

Lowest observed potency 2.4 1.4 0.04 0.07 A–4

Estimated dose associated with 10% increased incidence, in mg/kg-d.

95% lower bound on ED10, in mg/kg-d.

Per mg/kg-d, computed as 0.10/ED10.

Per mg/kg-d, computed as 0.10/LED10.

Bound from Norback and Weltman (1985).

These ranges reflect experimental uncertainty and variability of commercial

mixtures, but not human heterogeneity and differences between commercial and

environmental mixtures. Environmental processes have profound effects that can

increase or decrease toxicity, so toxicity of an environmental mixture is only partly

determined by the original commercial mixture. Potency estimates for an Aroclor tested

in the laboratory may not be the best surrogate for assessing that Aroclor as altered in

the environment. Sections 4 and 5 develop specific guidance for applying these

potency ranges to environmental mixtures.

The new upper-bound slopes are lower than the previous estimate of 7.7 per

mg/kg-d average lifetime exposure (U.S. EPA, 1988a). The previous estimate was

derived from female rats in the Norback and Weltman (1985) study; the new estimate

from the same study is 2.2 per mg/kg-d. This difference is attributable to three factors,

each responsible for reducing the slope by approximately one-third: the rat liver tumor

reevaluation (Moore et al., 1994), use of the new cross-species scaling factor (U.S.

EPA, 1992b), and not using a time-weighted average dose (see previous footnote).

The difference between the highest observed new upper-bound slope (2.2 per mg/kg-d)

and the lowest (0.07 per mg/kg-d) is entirely attributable to the available of tests on

several commercial mixtures (Brunner et al., 1996). This 30-fold range in potency

reflects differences in commercial mixture composition.

The different responses for male and female rats (Brunner et al., 1996) suggest

the possibility of developing different potency values for males and females. In view of

the 91 percent response in male Wistar rats (Schaeffer et al., 1984), as well as the

sensitivity of male mice (Kimbrough and Linder, 1974; Ito et al., 1973), it is premature

to conclude that females are always more sensitive.

For the thyroid tumors, no meaningful ED10 or LED10 can be computed. The

dose-response curves for each Aroclor are virtually horizontal across the experimental

range, so mathematical models cannot determine whether a dose-response curve

begins to be sublinear immediately below the experimental range or whether it remains

horizontal for several orders of magnitude below the experimental range before

becoming sublinear. This difficulty transcends PCBs and thyroid tumors; in general,

there would be an unacceptable level of uncertainty in using a sublinear extrapolation

approach when study results in the experimental range are not at all sublinear.

3.4. ANALYSES OF CONGENER TOXICITY

McFarland and Clarke (1989) explain how toxicity of some PCB congeners is

correlated with induction of mixed-function oxidases. Some congeners are described

as phenobarbital-type inducers, others as 3-methylcholanthrene-type inducers, and

some as having mixed inducing properties. The latter two groups most resemble

2,3,7,8-tetrachlorodibenzo-p-dioxin in structure and toxicity. Based on potential for

toxicity (some forms of toxicity, for example, neurotoxicity, may not be well represented)

and frequency of occurrence in environmental samples, 36 congeners of highest

concern were identified and classified (see table 3–3).

Table 3–3. PCB congeners of highest concern

Highest toxicity High toxicity Abundant in Potential

and abundance

and abundance environment for toxicity

ab c d

3–MC-type inducers: PB-type inducers: 18: 2,5,2'–TrCB 37: 3,4,4'–TrCB

77: 3,4,3',4'–TeCB 87: 2,3,4,2',5'–PeCB 44: 2,3,2',5'–TeCB 81: 3,4,5,4'–TeCB

126: 3,4,5,3',4'–PeCB 99: 2,4,5,2',4'–PeCB 49: 2,4,2',5'–TeCB 114: 2,3,4,5,4'–PeCB

169: 3,4,5,3',4',5'–HxCB 101: 2,4,5,2',5'–PeCB 52: 2,5,2',5'–TeCB 119: 2,4,6,3',4'–PeCB

153: 2,4,5,2',4',5'–HxCB 70: 2,5,3',4'–TeCB 123: 3,4,5,2',4'–PeCB

Mixed-type inducers: 180: 2,3,4,5,2',4',5'–HpCB 74: 2,4,5,4'–TeCB 157: 2,3,4,3',4',5'–HxCB

105: 2,3,4,3',4'–PeCB 183: 2,3,4,6,2',4',5'–HpCB 151: 2,3,5,6,2',5'–HxCB 158: 2,3,4,3',4',6'–HxCB

118: 2,4,5,3',4'–PeCB 194: 2,3,4,5,2',3',4',5'–OCB 177: 2,3,5,6,2',3',4'–HpCB 167: 2,4,5,3',4',5'–HxCB

128: 2,3,4,2',3',4'–HxCB 187: 2,3,5,6,2',4',5'–HpCB 168: 2,4,6,3',4',5'–HxCB

138: 2,3,4,2',4',5'–HxCB 201: 2,3,4,5,2',3',5',6'–OCB 189: 2,3,4,5,3',4',5'–HpCB

156: 2,3,4,5,3',4'–HxCB

170: 2,3,4,5,2',3',4'–HpCB

Pure 3-methylcholanthrene-type inducers and mixed-type inducers reported frequently in environmental samples.

Phenobarbital-type inducers reported frequently in environmental samples.

Weak inducers or noninducers reported frequently in environmental samples.

Mixed-type inducers not reported frequently in environmental samples, but toxicologically active.

Source: Adapted from McFarland and Clarke (1989).

U.S. EPA (1991) examined toxic effects, including cancer, of four structural

classes: dioxin-like PCBs, ortho-substituted PCBs, hydroxylated metabolites, and

sulfonated metabolites. Different mechanisms were discussed for dioxin-like and other

PCBs. It was concluded that congener toxicity could not be characterized by chlorine

content alone. Before adopting toxic equivalence factors (TEFs) for PCB congeners, it

was recommended to define other classes of PCBs and identify the mechanisms

involved. Criteria for developing TEFs were listed as (1) a demonstrated need, (2) a

well defined group of chemicals, (3) a broad base of toxicological data, (4) consistency

in the relative toxicity of congeners across toxicological endpoints, (5) demonstrated

additivity between the toxicity of individual congeners, (6) a mechanistic rationale, and

(7) consensus.

Safe (1990, 1994) characterized dioxin-like PCBs as eliciting a spectrum of

biochemical and toxic responses similar to chlorinated dibenzo-p-dioxins and

dibenzofurans, all acting through the aryl hydrocarbon receptor. Based on quantitative

structure-activity studies, the first conservative TEFs for dioxin-like PCBs were

proposed and refined. Use of these TEFs is limited to responses mediated through the

aryl hydrocarbon receptor.

Subsequently, WHO derived TEFs for dioxin-like PCBs (Ahlborg et al., 1994).

Included were congeners that show structural similarity to chlorinated

dibenzo-p-dioxins and dibenzofurans, bind to the aryl hydrocarbon receptor, elicit

dioxin-specific biochemical and toxic responses, and persist and accumulate in the

food chain. On the basis of these criteria, 13 PCB congeners were assigned TEFs,

expressed as a fraction of the toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (see

table 3–4). As new information is developed, these TEFs could change, or TEFs for

additional congeners could be developed. Section 4 gives guidance, and section 5 an

example, for applying these TEFs where a congener analysis is available.

Table 3–4. WHO interim TEFs for human intake of dioxin-like PCBs

Non-ortho congeners TEF Mono-ortho congeners TEF Di-ortho congeners TEF

77: 3,4,3',4'–TeCB 0.0005 105: 2,3,4,3',4'–PeCB 0.0001 170: 2,3,4,5,2',3',4'–HpCB 0.0001

126: 3,4,5,3',4'–PeCB 0.1 114: 2,3,4,5,4'–PeCB 0.0005 180: 2,3,4,5,2',4',5'–HpCB 0.00001

169: 3,4,5,3',4',5'–HxCB 0.01 118: 2,4,5,3',4'–PeCB 0.0001

123: 3,4,5,2',4'–PeCB 0.0001

156: 2,3,4,5,3',4'–HxCB 0.0005

157: 2,3,4,3',4',5'–HxCB 0.0005

167: 2,4,5,3',4',5'–HxCB 0.00001

189: 2,3,4,5,3',4',5'–HpCB 0.0001

Source: Adapted from Ahlborg et al. (1994).

Brown (1994) calculated relative metabolic rates for 146 congeners and

combined them into a measure of "relative human accumulability" for each Aroclor,

based on congener concentrations in the Aroclors. Assuming a correlation between

relative human accumulability and chronic toxicity or cancer, it was suggested to

calculate relative human accumulability (relative to Aroclor 1260) for each

environmental mixture. Because environmental processes can alter the relative

congener concentrations, this approach can yield cancer slopes outside the range

determined previously for commercial mixtures. In premeeting comments on this

assessment (U.S. EPA, 1996b), Brown refined this approach to express cancer slope

as a function of both relative human accumulability and dioxin toxic equivalents.

Congener concentrations for the commercial mixtures in the Brunner et al. (1996) study

will be used to evaluate this suggested approach.

4. APPLICATION OF THE DOSE-RESPONSE ASSESSMENT

4.1. APPLICATION TO PCB MIXTURES IN THE ENVIRONMENT

After release into the environment, PCB mixtures change through partitioning,

transformation, and bioaccumulation, differing considerably from commercial mixtures.

How can toxicity values for commercial mixtures be applied to mixtures in the

environment?

A consensus is emerging on the difficulty of assessing environmental mixtures

by reference to Aroclors. Safe (1994) wrote, "Regulatory agencies and environmental

scientists have recognized that the composition of PCBs in most environmental extracts

does not resemble the composition of the commercial products." Along similar lines,

ATSDR (1993) advised,

It is important to recognize that the PCBs to which people may be

exposed are likely to be different from the original PCB source because of

changes in congener and impurity composition resulting from differential

partitioning and transformation in the environment and differential

biological metabolism and retention. Because of this concern, current

data are considered inadequate to differentiate between the toxicity and

carcinogenicity of environmental PCB mixtures with any reasonable

degree of confidence.

For these reasons, risks from environmental mixtures are not assessed by

characterizing environmental mixtures as if they were Aroclors. This does not mean

that all environmental mixtures are regarded as equally potent; environmental mixtures

differ from commercial mixtures and from each other. To make distinctions about risks

from environmental mixtures, the range of potency observed for commercial mixtures

can be considered along with factors that increase or decrease risk.

First among these is persistence and bioaccumulation through the food chain.

Each species, in turn, retains persistent congeners that prove resistant to metabolism

and elimination (Oliver and Niimi, 1988). Bioaccumulated PCBs appear to be more

toxic than commercial PCBs (Aulerich et al., 1986). Mink fed Great Lakes fish

contaminated with PCBs showed liver and reproductive toxicity comparable to mink fed

Aroclor 1254 at quantities three times greater (Hornshaw et al., 1983). It is crucial to

recognize that commercial PCBs tested in laboratory animals were not subject to prior

selective retention of persistent congeners through the food chain. For exposure

through the food chain, risks can be higher than those estimated in this assessment.

Also important is the presence or absence of congeners and metabolites that

contribute to cancer induction. Mechanistic studies are beginning to identify these

congeners and describe their modes of action. Several congeners have dioxin-like

activity (Ahlborg et al., 1994; Safe, 1994), some promote tumors through different

modes of action (Buchmann et al., 1991). Because concentrations of these congeners

are altered by partitioning, transformation, and bioaccumulation in the environment,

risks from environmental mixtures can differ from those of commercial mixtures.

Congener analyses of environmental samples can provide information on the extent of

this difference and can be an important tool in risk assessment, particularly when fish

consumption is an issue.

Chlorine content was formerly regarded by some scientists as correlated with

cancer risk. Recently, however, Aroclor 1254 was found to be more potent than 1260,

which was only slightly more potent than 1242 (Brunner et al., 1996). This casts doubt

on chlorine content being a useful indicator of cancer potency in this range of chlorine

content; both the number and position of chlorines are important. It is instructive to

compare how the Aroclors rank by other measures. With respect to resistance to

metabolism and persistence in the body, there is an association with chlorine content,

which partially explains the greater experimental potency of commercial mixtures with

higher chlorine content. With respect to dioxin toxic equivalents (TEQs), however,

several studies have ranked Aroclor 1254, 1248, and 1242 as more potent than 1260

(Harper et al., 1995; Safe, 1994; Harris et al., 1993; Hong et al., 1993; Schulz et al.,

1989). The combined effect is difficult to predict, as Aroclor 1260 and mixtures with

higher chlorine content have lower dioxin TEQs but persist longer in the environment

and in the body.

A key finding is the several-fold lower potency of Aroclor 1016 compared with

1242 (Brunner et al., 1996). Though these mixtures are similar in average chlorine

content (41 and 42 percent, respectively), Aroclor 1016 has virtually no congeners with

more than four chlorines. This suggests that one way to differentiate less potent

mixtures is to verify the absence of congeners with more than four chlorines.

Since mixtures of congeners with more than four chlorines cannot be assessed

using chlorine content alone, other determinants of toxicity are used. Two important

determinants, persistence and bioaccumulation, can be related to exposure pathway.

(Persistence is not synonymous with toxicity; however, in the absence of testing on

most congeners, it is reasonable to suppose some correlation between persistence and

toxicity.) Evaporated or dissolved congeners tend to be lower in chlorine content than

the original mixture; they tend also to be more inclined to metabolism and elimination

and lower in persistence and toxicity. On the other hand, congeners adsorbed to

sediment or soil tend to be higher in chlorine content and persistence, and

bioaccumulated congeners ingested through the food chain tend to be highest of all.

Rates of these processes vary over several orders of magnitude (Hutzinger et al.,

1974; Erickson, 1986), thus the effect of environmental processes can be greater than

the spread in potency or slope estimated from commercial mixtures.

For these reasons, a tiered approach is recommended. The default tier uses

exposure pathway to choose appropriate potency values from the ranges described in

table 3–2. The highest observed potency from these ranges is appropriate for food

chain exposure, sediment or soil ingestion, and dust or aerosol inhalation, pathways

where environmental processes tend to increase risk. Lower potencies are appropriate

for ingestion of water-soluble congeners or inhalation of evaporated congeners,

pathways where environmental processes tend to decrease risk. To the extent that

drinking water or ambient air contains contaminated sediment or dust, the higher

potency values would be appropriate, as congeners adsorbed to sediment or dust tend

to be of high chlorine content and persistence, especially for sediment or dust with high

organic content. Since the lowest observed potency, based on studies to date, is

derived from Aroclor 1016, its use is most appropriate in the absence of congeners with

more than four chlorines. When congeners with more than four chlorines are present,

but exposure is by drinking water ingestion or vapor inhalation, potency values derived

from the next lowest tested mixture, Aroclor 1242, can be more appropriate.

Thus three reference points can be designated for each range (see table 4–1):

a "high risk" point based on studies of Aroclor 1260 and 1254, which give the highest

observed potencies; a "low risk" point, based on the study of Aroclor 1242; and a

"lowest risk" point, based on the study of Aroclor 1016. The "high risk" point is used for

exposure pathways associated with environmental processes that tend to increase risk;

the "low risk" point, for exposure pathways that tend to decrease risk; and the "lowest

risk" point, for cases where congener or isomer analyses verify that congeners with

more than four chlorines comprise less than one-half percent of total PCBs, suggesting

that potency is best represented by the least potent tested mixture. This demonstrates

a potential use of congener analysis of environmental samples. Section 5 provides a

series of examples illustrating this tiered approach.

Table 4–1. Tiers of human potency and slope estimates for environmental mixtures

HIGH RISK AND PERSISTENCE

Central Upper-bound

ED10

LED10 slope slope Criteria for use

abcd

0.086 0.067 1. 2. Food chain exposure

Sediment or soil ingestion

Dust or aerosol inhalation

Dermal exposure, if an absorption factor has

been applied to reduce the external dose

Presence of dioxin-like, tumor-promoting, or

persistent congeners in other media

Early-life exposure (all pathways and mixtures)

LOW RISK AND PERSISTENCE

Central Upper-bound

ED10

LED10 slope slope Criteria for use

abcd

0.38 0.27 0.3 0.4 Ingestion of water-soluble congeners

Inhalation of evaporated congeners

Dermal exposure, if no absorption factor has

been applied to reduce the external dose

LOWEST RISK AND PERSISTENCE

Central Upper-bound

ED10

LED10 slope slope Criteria for use

abcd

2.4 1.4 0.04 0.07 Congener or isomer analyses verify that

congeners with more than 4 chlorines comprise

less than 1/2% of total PCBs

Estimated dose associated with 10% increased incidence, in mg/kg-d.

95% lower bound on ED10, in mg/kg-d.

Per mg/kg-d, computed as 0.10/ED10 and rounded to one significant digit.

Per mg/kg-d, computed as 0.10/LED10 and rounded to one significant digit.

This reasoning assumes that the PCB mixture has undergone alteration in the

environment for many years and that partitioning into different environmental media has

approached equilibrium. Congener or isomer analyses are important to verifying

equilibrium. For example, if drinking water samples contain high concentrations of

congeners with high chlorine content, this could indicate a recent release of PCBs that

has not had sufficient time to partition as expected, a continuing release of PCBs, or

the presence of stirred-up sediment and the subsequent release of adsorbed

congeners with high chlorine content. Judgment should be used in assessing these

situations, which are most likely not best represented by the "lowest risk" point derived

from Aroclor 1016.

4.2. APPLICATION TO DIFFERENT ROUTES OF EXPOSURE

What inferences can be made about dermal or inhalation exposures, two routes

for which there are no lifetime cancer studies?

PCBs are absorbed through ingestion, inhalation, and dermal exposure, after

which they are transported similarly through the circulation. This provides a

reasonable basis for expecting similar internal effects from different routes of

environmental exposure. In addition, the capacity of Aroclor 1254 to promote mouse

lung tumors (Anderson et al., 1983, 1994; Beebe et al., 1992, 1993) and the

observation of skin cancer from occupational exposure (Sinks et al. 1992) suggest a

potential for cancer at the point of entry.

Dermal absorption through human skin is rapid initially and continues at a slower

rate with further contact (Wester et al., 1987, 1990). Rhesus monkeys exposed to

Aroclor 1242 or 1254 in soil for 24 hours absorbed 14 percent through the skin (Wester

et al., 1993). Absorption of PCBs from soil involves competition between the lipophilic

attraction of PCBs to skin and adsorption to organic soil material. Absorption increases

with duration of skin contact and moisture content of soil, but these relationships have

not been quantified. Overall, use of the low end of the potency ranges for dermal

exposure appears appropriate in light of the substantial but incomplete absorption

through the skin. If, however, an estimate of dermal exposure has been reduced by an

absorption factor, then the high end of the potency ranges would be appropriate, as the

use of an absorption factor converts an external dose to an internal dose.

The lung is also a potential target for cancer risk following PCB exposure.

Aroclor 1254 is active as a promoter in the mouse lung initiated with methylating

nitrosamines. Relevant mechanistic information includes: (1) persistent induction of

cytochrome P–450 1A in the lung (up to several months after a single PCB dose),

(2) promotion by 2,3,4,2',4',5'–HxCB, not by 2,4,5,2',4',5'–HxCB, and partial abrogation

of 2,3,4,2',4',5'–HxCB's effects by 2,4,5,2',4',5'–HxCB, (3) correlation of the promotion

effect with body burden of 2,4,5,2',4'–PeCB, and (4) selective retention of PCB

congeners, especially 2,3,4,3',4'–PeCB, in mouse lung (Anderson, 1991b).

Inhaled PCBs can be rapidly absorbed, although rates have not been quantified

(ATSDR, 1993). Rapid absorption, however, suggests potency by inhalation is

comparable to potency by ingestion. Because PCBs are slowly metabolized, little

uncertainty results from the first-pass effect, where ingested toxicants are subject to

metabolism in the liver before entering the circulation, while inhaled toxicants enter the

circulation before reaching the liver. As with ingested mixtures, the composition of an

inhaled mixture influences its toxicity. Evaporated mixtures tend to have low chlorine

content and persistence, mixtures adsorbed to dust and soil tend to be high in this

regard, and mixtures suspended in an aerosol can be more diverse.

4.3. APPLICATION TO LESS-THAN-LIFETIME AND EARLY-LIFE EXPOSURES

In assessing cancer risks from less-than-lifetime exposure, the common practice

is to prorate cumulative exposure over the lifespan (U.S. EPA, 1986a, 1992a). For

example, exposure lasting 7 years of a 70-year lifespan would be assumed to have

one-tenth the effect of lifetime exposure. Does the information available for PCBs

support this default or suggest an alternative?

Less-than-lifetime exposure induced statistically significant increased incidences

of liver tumors in female rats fed Aroclors 1260, 1254, and 1242 (Brunner et al., 1996).

This result was most pronounced for Aroclor 1260, where tumor incidences at the

highest dose were higher for a 12-month exposure than for a 24-month lifetime

exposure. Only Aroclor 1016 showed no significant increases from less-than-lifetime

exposure. The earlier less-than-lifetime studies in rats and mice suggest that less-

than-lifetime exposure can quickly induce high incidences of early stages of tumor

development (Kimbrough et al., 1972; Ito et al., 1973, 1974). With further exposure,

these can progress to malignancy (Kimbrough et al., 1975; Norback and Weltman,

1985). Tumor incidences from less-than-lifetime exposure were sometimes lower

(Kimbrough and Linder, 1974), and sometimes similar (Rao and Banerji, 1988), to those

from full lifetime exposure.

The strong response for Aroclor 1260, coupled with the lack of response for

Aroclor 1016, suggests that these findings may be related to persistence in the body.

PCBs entering the body are transported by the circulation to internal organs and fat,

where they are stored (Matthews and Anderson, 1975). Equilibrium is maintained

among external exposure levels, concentrations in blood, and concentrations in fat and

other tissues. When external exposure is reduced, to maintain equilibrium, stored

PCBs reenter the circulation and provide a continuing internal source of exposure

(Matthews and Anderson, 1975). Thus PCBs from short-term exposure can be stored

in the body and emerge as a source of exposure much later.

Persistence in the body can enhance the opportunity for PCB congeners to

express tumor promoting activity (Safe, 1994). Persistent congeners can retain

biological activity long after exposure stops (Anderson et al., 1991a); some persistent

congeners are tumor promoters. The congener 3,4,3',4'-tetrachlorobiphenyl continues

to promote tumors vigorously after dosing stops, but not 2,4,2',5'-tetrachlorobiphenyl

(Buchmann et al., 1991; Luebeck et al., 1991). Although the probability of liver focus

extinction is high after dosing stops, those that become large tend to persist (Luebeck

et al., 1991). This would allay concern for short-term exposure but increase concern as

exposure duration increases. Further studies of less-than-lifetime exposure, as well as

methods for quantifying the differential effects of less-than-lifetime exposure, are

needed.

Regarding early-life exposure, infants can be highly exposed to PCBs during

pregnancy and lactation (Dewailly et al., 1991, 1994). The accumulation of PCBs in

human adipose tissue creates a store for subsequent release of PCBs into the

bloodstream and then into the fetal circulation. During the postpartum period, PCBs

are mobilized from adipose stores, transferred into human milk, and delivered to the

neonate via nursing (Dewailly et al., 1991). This source of exposure may account for a

substantial fraction of chlorinated dibenzo-p-dioxins and dibenzofurans, and the same

may be true for dioxin-like and other PCBs. It is, therefore, important to assess the

extent of exposure through the human milk pathway; if direct measurement of

concentrations in milk are not available, estimates can be derived from maternal

exposures (Smith, 1987).

Normal fetal development depends on the timing and rate of release of T3 and

T4. Some evidence indicates that PCBs can alter normal T3 and T4 metabolism,

thereby disturbing thyroid function and provoking secondary impacts on organogenesis

during development. Any estrogenic/anti-estrogenic, androgenic/anti-androgenic, or

other hormonal activity of PCB mixtures has the possibility of altering the development

of reproductive organs or the urogenital tract, potentially causing cancer or other

adverse effects through a mechanism different from those causing liver cancer (U.S.

EPA, 1996b).

Few studies, however, have investigated early-life sensitivity. In human infants,

glucuronidative mechanisms are not fully developed; additionally, some nursing infants

receive a steroid in human milk that inhibits the activity of glucuronyl transferase,

further reducing PCB metabolism and elimination (Calabrese and Sorenson, 1977). In

animals, Aroclor 1260 induced high incidences of liver tumors when fed to 5-week-old

rats for a short time (Rao and Banerji, 1988). On the other hand, acute perinatal

dosing with Aroclor 1254 promoted nitrosamine-initiated lung and liver tumors in mice

but did not induce cancer in the offspring when administered alone (Anderson et al.,

1983, 1986, 1994). A study of polybrominated biphenyls (PBBs) found that perinatal

exposure enhanced susceptibility to liver tumors for female rats also exposed as adults

and increased the incidence of liver tumors in male and female mice not further

exposed as adults (NTP, 1993). Because of the potential magnitude of early-life

exposures, the possibility of greater perinatal sensitivity, and the likelihood of

interactions among thyroid and hormonal development, it is reasonable to conclude

that early-life exposures may be associated with increased risks; this would indicate

using the "high-risk" potency estimates for early-life exposure. A method for

quantifying the differential effects of early-life exposure is needed.

4.4. APPLICATION WITH DIOXIN TOXIC EQUIVALENCE FACTORS

TEFs have been proposed for some dioxin-like PCB congeners, while this

assessment develops potency values for overall concentrations of mixtures. How can

the TEF approach supplement the mixture-based approach?

When assessing PCB mixtures, it is important to recognize that both dioxin-like

and nondioxin-like modes of action contribute to overall PCB toxicity (Safe, 1994;

McFarland and Clarke, 1989; Birnbaum and DeVito, in press). Because relatively few

PCB congeners are dioxin-like, dioxin equivalence explains only part of a PCB

mixture's toxicity. (This applies to cancer and other forms of toxicity, for example,

neurotoxicity and endocrine disruption; Birnbaum and DeVito, in press.) Hence, PCB

assessments should begin with the mixture-based approach developed in this report.

At the same time, it is possible that concentrations of dioxin-like congeners are

increased in an environmental mixture. When congener concentrations are available,

the mixture-based approach can be supplemented by analysis of dioxin TEQs to

evaluate dioxin-like toxicity. Section 5 gives an example for calculating dioxin TEQs

when a congener analysis is available.

When assessing mixtures of dioxin and related compounds, it is important to

consider the contribution of dioxin-like PCBs to total dioxin equivalents (U.S. EPA,

1994b). TEQs for dioxin-like PCBs (Ahlborg et al., 1994) can be added to those for

other dioxin-like compounds. In some situations, PCBs can contribute more dioxin-like

toxicity than chlorinated dibenzo-p-dioxins and dibenzofurans (Schecter et al., 1994;

Dewailly et al., 1991, 1994). The congener 2,4,5,3',4'-pentachlorobiphenyl, shown to

have tumor-promoting activity, is a major contributor to total dioxin equivalents in the

United States (Patterson et al., 1994) and maritime Quebec (Dewailly et al., 1994).

Exposure to dioxin-like PCBs adds to background exposure of dioxin-like

compounds and augments processes associated with dioxin toxicity. There is support

for using low-dose-linear dose-response models for incremental doses that add to

existing background exposure (Crump et al., 1976; Lutz, 1990). Thus confidence in

this assessment's use of low-dose-linear models is enhanced for the dioxin-like portion

of a PCB mixture.

5. CHARACTERIZATION AND GUIDANCE FOR RISK ASSESSORS

5.1. DOSE-RESPONSE CHARACTERIZATION

Joint consideration of cancer studies and environmental processes leads to a

conclusion that environmental PCB mixtures are highly likely to pose a risk of cancer to

humans. Although environmental mixtures have not been tested in cancer assays, this

conclusion is supported by several complementary sources of information. Statistically

significant, dose-related, increased incidences of liver tumors were induced in female

rats by Aroclors 1260, 1254, 1242, and 1016 (Brunner et al., 1996). These mixtures

contain overlapping groups of congeners that, together, span the range of congeners

most frequently found in environmental mixtures. Several congeners promote tumors

or have dioxin-like activity; these congeners are found in environmental samples and in

a variety of organisms, including humans.

The range of potency observed for commercial mixtures is used to represent the

potency of environmental mixtures. The range reflects experimental uncertainty and

variability of commercial mixtures, but not human heterogeneity or differences between

commercial and environmental mixtures. Environmental processes alter mixtures

through partitioning, transformation, and bioaccumulation, thereby decreasing or

increasing toxicity. The overall effect can be considerable, and the range observed for

commercial mixtures may underestimate the true range for environmental mixtures.

Limiting the potency of environmental mixtures to the range observed for commercial

mixtures reflects a decision to base potency estimates on experimental results,

however uncertain, rather than apply safety factors to compensate for lack of

information (U.S. EPA, 1996b).

A tiered approach allows use of different kinds of information in estimating the

potency of environmental mixtures. When congener information is limited, exposure

pathway is used to indicate whether environmental processes have decreased or

increased a mixture's potency. Partitioning, transformation, and bioaccumulation have

been extensively studied and can be associated with exposure pathway, thus the use

of exposure pathway to represent environmental processes increases confidence in the

risks inferred for environmental mixtures. When available, congener information is an

important tool for refining a potency estimate that was based on exposure pathway.

Extrapolation to environmental levels is based on models that are linear at low

doses. Low-dose-linear models are appropriate when a carcinogen acts in concert with

other exposures and processes that cause a background incidence of cancer. Even

when the mode of action indicates a nonlinear dose-response curve in homogeneous

animal populations, the presence of genetic and lifestyle factors in a heterogeneous

human population tends to make the dose-response curve more linear.

Depending on the specific application, either central estimates or upper bounds

can be appropriate. Central estimates describe a typical individual's risk, while upper

bounds provide assurance that this risk is not likely to be underestimated if the

underlying model is correct. The upper bounds calculated in this assessment reflect

study design and provide no information about sensitive individuals or groups. Central

estimates are useful for estimating aggregate risk across a population. Central

estimates are used for comparing or ranking environmental hazards, while upper

bounds provide information about the precision of the comparison or ranking.

Comparing a central estimate with its upper bound indicates whether the central

estimate is stable enough to support credible risk estimates. In this assessment, the

less-than-twofold difference between central estimates and upper bounds indicates that

these estimates are stable.

Uncertainty around these estimates extends in both directions. The slope factor

ranges primarily reflect mixture variability, and so are not necessarily appropriate for

probabilistic analyses that attempt to describe model uncertainty and parameter

uncertainty. Several sources of uncertainty are inherent in the experimental

information used in this assessment:

Experimental design and conduct: The new rat study (Brunner et al., 1996) is

quite extensive in design and conduct, going beyond standard designs for

cancer studies in many respects.

Variability in commercial mixture composition: For the four Aroclors tested

in female Sprague-Dawley rats (Brunner et al., 1996; Norback and Weltman,

1985), there is a 30-fold range in potency. This whole range is used to

represent environmental mixtures.

Variability across strains: In the four rat strains tested, sensitivity varies up to

15-fold. Potency and slope estimates were derived from a strain covering the

middle of this range.

Variability between sexes: Potency and slope estimates were derived from

female rats, whose liver response was usually greater than that of males. The

greatest response in the liver, however, was in male rats. Greater sensitivity of

females was not seen in mice, nor in the thyroid.

Variability across experiments: For the same Aroclor, sex, and strain,

differences up to four-fold were observed. To reflect this lot-to-lot variability,

both estimates were included.

Experimental uncertainty (sample size): Central and upper-bound potency

estimates differ by no more than about two-fold. This is a minor source of

uncertainty.

Potential for other carcinogenic effects: The new rat study reported small

increases in thyroid tumor incidence for male rats, suggesting a potential for a

hormonal mode of action. These results have not yet been publicly discussed or

peer reviewed.

Other sources of uncertainty arise in the methods for assessing this

experimental information and applying it to human environmental exposure:

Animal-to-human extrapolation: The use of default cross-species scaling

factors is intended as an unbiased projection not expected to provide

conservatism (U.S. EPA, 1992b). Information is lacking to evaluate whether

humans are more or less sensitive than rats.

High-to-low-dose extrapolation: The use of models that are linear at low

doses can potentially overestimate potency by an unknown amount. The rat

studies, however, show no evidence of sublinearity in the experimental range.

Route-to-route extrapolation: Information on relative absorption rates

suggests that differences in toxicity across exposure routes are small.

Difference between commercial and environmental mixtures: Commercial

mixtures released into the environment are altered by environmental processes.

Qualitatively, exposure pathway is a reasonably good indicator of whether

potency has been decreased or increased. Quantitatively, the percentage

change in toxicity is unknown, though the 30-fold range in potency observed for

commercial mixtures likely underestimates the range for environmental mixtures.

Persistence and exposure duration: Some PCBs persist in the body and

retain biological activity after exposure stops (Anderson et al., 1991a).

Compared with the current default practice of assuming that less-than-lifetime

effects are proportional to exposure duration, rats exposed to the persistent

mixture Aroclor 1260 had more tumors, while rats exposed to the less persistent

Aroclor 1016 had fewer tumors (Brunner et al., 1996). Thus the current default

practice can underestimate risks for persistent mixtures.

Human variability in sensitivity: People with decreased liver function can

have less capacity to metabolize and eliminate PCBs. Approximately five

percent of nursing infants receive a steroid in human milk that further inhibits

PCB metabolism and elimination (Calabrese and Sorenson, 1977).

Human variability in exposure: Blood concentrations vary over a 100-fold

range (ATSDR, 1993). Highly exposed populations include nursing infants,

consumers of game animals contaminated through the food chain, and workers

with occupational exposure. There is greater confidence in risk estimates for

highly exposed groups.

When exposure involves the food chain, uncertainty extends principally in one

direction: through the food chain, living organisms selectively bioaccumulate persistent

congeners, but commercial mixtures tested in laboratory animals were not subject to

prior selective retention of persistent congeners. Bioaccumulated PCBs appear to be

more toxic than commercial PCBs (Aulerich et al., 1986; Hornshaw et al., 1983) and

appear to be more persistent in the body (Hovinga et al., 1992). For exposure through

the food chain, risks can be higher than those estimated in this assessment. Two

highly exposed populations, nursing infants and consumers of contaminated game

animals, are exposed through food.

The dioxin-like nature of some PCBs raises a concern for cumulative exposure,

as dioxin-like congeners add to background exposure of other dioxin-like compounds

and augment processes associated with dioxin toxicity. This weighs against

considering PCB exposure in isolation or as an increment to a background exposure of

zero. Confidence in this assessment's use of low-dose-linear models is enhanced

when there is additivity to background exposures and processes.

To gauge the distance of an extrapolation, human exposures can be compared

with the LED10. For mixtures that are altered in the environment, however, this simple

comparison of relative exposure is of limited value as some congeners increase in

concentration while others decrease. Additionally, for exposures that augment

processes leading to a background incidence of cancer, identifying and quantifying

background human exposures are necessary for a true measure of total exposure.

5.2. INFLUENCE OF PROPOSED CANCER GUIDELINES

This assessment demonstrates several ideas from EPA's proposed cancer

guidelines (U.S. EPA, 1996a). Most prominent is development of a range of potency

estimates, using studies for a range of mixtures, instead of focusing on the highest-

potency mixture. For low-dose extrapolation, an LED10 approach replaces the

linearized multistage procedure. An ED10 approach provides a statistically stable

method for deriving central estimates of low-dose slopes. Dose calculations use the

interagency consensus cross-species scaling factor, based on the 3/4 power of relative

body weight (U.S. EPA, 1992b).

Also evident is the proposed guidelines' encouragement to use different kinds of

information. It is reflected in the tiered approach that uses site-specific congener or

isomer analyses when available, but can differentiate among environmental mixtures

using exposure pathway when mixture information is limited. It is also found in this

assessment itself, which combines information on toxicity and environmental

processes.

The proposed guidelines' emphasis on discussing circumstances that affect

cancer risks, especially exposure route considerations, is found throughout this

assessment. There is extensive discussion of how environmental processes alter the

composition and toxicity of PCB mixtures. Exposure circumstances are addressed in a

framework that distinguishes different exposure pathways as lower risk or higher risk.

None of these features, however, is inconsistent with previous guidelines (U.S.

EPA, 1986a), whose intent is "to permit sufficient flexibility to accommodate new

knowledge and new assessment methods as they emerge." Each new feature of this

assessment can be viewed in this spirit.

5.3. RESEARCH NEEDS

This dose-response assessment has tried to make the best use of the available

information. For some questions it has stopped short because information or methods

are needed to resolve an important issue. Research that would enable this

assessment to proceed further includes:

Cancer studies comparing commercial and environmental mixtures, especially

those found in the food chain. This assessment warns that food chain risks can

be underestimated, but the extent is not quantified.

A method for using lifetime studies to assess less-than-lifetime exposure to

persistent agents. This assessment warns that assuming risk and exposure

duration are proportional can underestimate risks for persistent mixtures, but the

extent is not quantified.

Relative cancer potency factors for congeners identified by McFarland and

Clarke (1989). This will improve evaluations of mixtures where carcinogenic

potential has been increased or decreased beyond the range of tested

commercial mixtures. To help determine such factors, 2-year studies in female

rats are recommended for 3,4,5,3',4'-pentachlorobiphenyl and 3,4,5,3',4',5'-

hexachlorobiphenyl, the two congeners with the highest dioxin TEFs, and

2,4,5,2',4',5'-hexachlorobiphenyl and 2,3,4,5,2',4',5'-heptachlorobiphenyl, the

two persistent congeners whose PBB analogues constitute over 80 percent of

the PBB mixture that caused hepatocellular adenomas, carcinomas, and

cholangiocarcinomas in rats.

Studies to test the hypothesis that the carcinogenic activity of Aroclor 1016 is

due to its tetrachlorobiphenyls only, with no contribution from congeners with

1–3 chlorines. Current information is inadequate to evaluate this hypothesis.

At the peer review workshop, the panel identified other areas where research

could help resolve important questions. These can be grouped broadly into exposure

methods research, effects research, and risk assessment methods development.

Research needs in exposure methods include:

Standard analytical methods, including sample preparation, for measuring PCB

congeners in environmental samples.

Database of congener levels in environmental samples.

Research needs on effects include:

Epidemiologic studies focused on tumor promotion. If PCBs act as tumor

promoters, they would increase cancer mainly in humans with already-initiated

cancer cells. For common cancers with complex etiologies, promotional effects

will be seen only if specifically looked for.

Mechanism-oriented dose-response data for environmental mixtures, including

promotional, hormonal, sex-specific effects.

Mechanisms of PCB-induced liver cancer in rats, its similarity to the

mechanism(s) of rat liver cancer induced by "nongenotoxic" carcinogens, its

activities at low dose levels, and its relevance to humans.

Identifying the most significant congeners in commercial and environmental

mixtures, describing their modes of action, and conducting studies to quantify

their slope factors.

Dose-response studies in a broader range of test animals.

Determine sensitivity of fetuses and newborns to the carcinogenic effects of

environmental mixtures.

Determine risks for thyroid and urogenital/reproductive tract cancers in

newborns and adults.

Methods development needs for risk assessment include:

Developing an appropriate dose metric for PCBs.

Exploring structure-activity methods for predicting pharmacokinetic parameters

for PCB mixtures.

Verifying the appropriateness, for PCBs, of the new cross-species scaling factor.

Evaluating the consistency of human and animal studies.

Developing quantitative uncertainty distributions for key sources of uncertainty.

Work in progress includes using congener compositions of the tested Aroclors

from the Brunner et al. (1996) study to evaluate the "relative human accumulability"

approach (Brown, 1994; U.S. EPA, 1996b). Congener compositions of the Aroclors

and the rat tissues can also be used in a factor analysis that would identify a subset of

congeners most associated with tumor induction. Field analyses can then reduce

uncertainty by quantifying a small number of critical congeners.

Elsewhere, new epidemiologic information is being analyzed. The National

Institute for Occupational Safety and Health is updating its study of the Indiana cohort,

and is expanding and updating its studies of the cohorts in Massachusetts and New

York. Additionally, the females in these three cohorts are being pooled for a study of

breast cancer.

Finally, future risk assessments of PCBs in the environment should consider

more than the risk of cancer. Although the purpose of this report was to evaluate

cancer risks, the emerging scientific literature indicates that toxicological endpoints

other than cancer may also be important to human health. These toxic effects should

be included along with cancer in future assessments of PCBs.

5.4. SUMMARY OF GUIDANCE FOR RISK ASSESSORS

Joint consideration of cancer studies and environmental processes leads to a

conclusion that environmental PCB mixtures are highly likely to pose a risk of cancer to

humans. The cancer potency of PCB mixtures is determined using a tiered approach

that depends on the information available (see table 4–1).

Upper-bound slope factors, derived by linear extrapolation from LED10s, are

described by a range of estimates with three reference points.

Upper-bound slope factors: 0.07 – 0.4 – 2 per mg/kg-d

Slope factors are multiplied by lifetime average exposure levels to estimate the risk of

cancer. The upper bounds reflect study design and provide no information about

sensitive individuals or groups. Although PCB exposures are often characterized in

terms of Aroclors, this can be both imprecise and inappropriate. Total PCBs or

congener or isomer analyses are recommended.

The first (default) tier is invoked when information on the mixture of interest is

limited. The upper reference point (2 per mg/kg-d) is appropriate for food chain

exposure, sediment or soil ingestion, and dust or aerosol inhalation; these are

exposure pathways for which environmental processes are likely to increase risk. Due

to potential for higher sensitivity early in life, the upper reference point is also used for

all early-life exposure. The middle reference point (0.4 per mg/kg-d) is appropriate for

drinking water ingestion and vapor inhalation; these are exposure pathways for which

environmental processes are likely to decrease risk. The lowest reference point (0.07

per mg/kg-d) should not be used without specific information on the congener

composition of the mixture.

The second tier is invoked when there are congener or isomer analyses for the

mixture of interest. The lowest reference point (0.07 per mg/kg-d) can be used if these

analyses verify that congeners with more than four chlorines comprise less than one-

half percent of total PCBs, as well as the absence of dioxin-like, tumor-promoting, and

persistent congeners. When congener concentrations are available, the slope-factor

approach can be supplemented by analysis of dioxin TEQs to evaluate dioxin-like

toxicity.

Central-estimate slope factors, derived by linear extrapolation from ED10s, can

be described by a similar range with three reference points.

Central-estimate slope factors: 0.04 – 0.3 – 1 per mg/kg-d

Central estimates describe a typical individual's risk, while upper bounds provide

assurance that this risk is not likely to be underestimated if the underlying model is

correct. Central estimates are useful for estimating aggregate risk across a population.

Highly exposed populations include nursing infants and consumers of game fish,

game animals, or products of animals contaminated through the food chain. Highly

sensitive populations include people with decreased liver function and infants.

A few limitations of this assessment should be noted:

It is crucial to recognize that commercial PCBs tested in laboratory animals were

not subject to prior selective retention of persistent congeners through the food

chain. Bioaccumulated PCBs appear to be more toxic than commercial PCBs

and appear to be more persistent in the body. For exposure through the food

chain, risks can be higher than those estimated in this assessment.

PCBs persist in the body, providing a continuing source of internal exposure

after external exposure stops. There may be greater-than-proportional effects

from less-than-lifetime exposure, especially for persistent mixtures and for early-

life exposures.

5.5. EXAMPLES

Example 1. Consider a release of PCBs onto the ground near a river or lake.

Potential pathways of human exposure include vapor inhalation, drinking water, fish

ingestion, and skin contact with ambient water and contaminated soil. The population

of interest includes anglers who consume an average of two 105-g portions of local fish

each week. They spend most of their time in the area, breathing 20 m air and drinking

2 L water, on average, each day. Skin contact with ambient water and soil is negligible

for this population. A 30-year exposure duration is to be considered, with a

representative lifespan of 70 years and body weight of 70 kg. Environmental samples

indicate long-term average concentrations of 0.01 g/m in ambient air, 5 g/L in

drinking water, and 110 g/kg in the edible portion of local fish. Dust in ambient air and

sediment in drinking water are negligible.

Because of partitioning, transformation, and bioaccumulation, different fractions

of the original mixture are encountered through these pathways, hence different

potency values are appropriate. Vapor inhalation is associated with "low risk" in

table 4–1 (evaporating congeners tend to have low chlorine content and be inclined to

metabolism and elimination), so the low end of the range (upper-bound slope of 0.4 per

mg/kg-d) is used for vapor inhalation. Similarly, ingestion of water-soluble congeners is

associated with "low risk" (dissolved congeners tend to have low chlorine content and

be inclined to metabolism and elimination), so the low end is also used for drinking

water. (If ambient air or drinking water had contained significant amounts of

contaminated dust or sediment, the high-end potency values would be appropriate, as

adsorbed congeners tend to be of high chlorine content and persistence.) Food chain

exposure appears is associated with "high risk" (aquatic organisms and fish selectively

accumulate congeners of high chlorine content and persistence that are resistant to

metabolism and elimination), so the high end of the range (upper-bound slope of 2 per

mg/kg-d) is used for fish ingestion.

The lifetime average daily dose (

LADD

) is calculated as the product of

concentration

, intake rate

, and exposure duration

divided by body weight

and lifetime

(U.S. EPA, 1992a)(table 5-1):

Table 5–1. Sample lifetime average daily dose calculations (examples 1 and 2)

Pathway

C IR ED BW LT LADD

Vapor inhalation 0.01 g/m 20 m /d 30 yr 70 kg 70 yr 1.2×10 mg/kg-d

33 –6

Drinking water 5. g/L 2 L/d 30 yr 70 kg 70 yr 6.1×10 mg/kg-d

–5

Fish ingestion 110 g/kg 30 g/d 30 yr 70 kg 70 yr 2.0×10 mg/kg-d

–5

LADD

/ (

)

For each pathway, the lifetime average daily dose is multiplied by the

appropriate slope to estimate risk (table 5-2):

Table 5–2. Sample risk calculations (example 1)

Pathway

LADD

Slope Risk

Vapor inhalation 1.2×10 mg/kg-d 0.4 per mg/kg-d 4.8×10

–6 –7

Drinking water 6.1×10 mg/kg-d 0.4 per mg/kg-d 2.4×10

–5 –5

Fish ingestion 2.0×10 mg/kg-d 2 per mg/kg-d 4.0×10

–5 –5

Sum 8.2×10 mg/kg-d 6.4×10

–5 –5

Risk =

LADD

× Slope

It is important to remember that this specific site exposure adds to a background

level of exposure from other sources.

Example 2. To show how additional, better information can improve the risk

assessment, suppose an analysis of PCB congeners in drinking water is performed in

the previous example. Suppose this analysis confirms that congeners with more than

four chlorines comprise less than a half-percent of total PCBs. Then it would be

plausible to treat PCBs in drinking water as similar to Aroclor 1016. The slope derived

from the Aroclor 1016 study could be used for this pathway(see table 5-3):

Table 5–3. Sample risk calculations (example 2)

Pathway

LADD

Slope Risk

Vapor inhalation 1.2×10 mg/kg-d 0.4 per mg/kg-d 4.8×10

–6 –7

Drinking water 6.1×10 mg/kg-d 0.07 per mg/kg-d 4.3×10

–5 –6

Fish ingestion 2.0×10 mg/kg-d 2 per mg/kg-d 4.0×10

–5 –5

Sum 8.2×10 mg/kg-d 4.5×10

–5 –5

Risk =

LADD

× Slope

The additional information leads to a conclusion that fish ingestion is the

principal pathway contributing to risk, and that drinking water and vapor inhalation are

of lesser consequence. It would be advisable to examine variability in fish consumption

rates and fish tissue concentrations to determine whether some individuals are at much

higher risk.

Example 3. Next, suppose an analysis of PCB congeners in the edible portion

of the fish is performed in the previous example, and it shows that concentrations of

dioxin-like congeners are greatly enhanced (see table 5-4):

Table 5–4. Sample congener concentrations and dioxin toxic equivalents (TEQs) in edible portion

of fish (example 3)

Congener

Conc. TEF TEQ

77: 3,4,3',4'–TeCB 2.1 g/kg 0.0005 0.0011 g/kg

105: 2,3,4,3',4'–PeCB 14. 0.0001 0.0014

114: 2,3,4,5,4'–PeCB 1.8 0.0005 0.0009

118: 2,4,5,3',4'–PeCB 54. 0.0001 0.0054

123: 3,4,5,2',4'–PeCB 1.4 0.0001 0.0001

126: 3,4,5,3',4'–PeCB 0.14 0.1 0.0140

156: 2,3,4,5,3',4'–HxCB 5.0 0.0005 0.0025

157: 2,3,4,3',4',5'–HxCB 1.1 0.0005 0.0006

167: 2,4,5,3',4',5'–HxCB 7.7 0.00001 0.0001

169: 3,4,5,3',4',5'–HxCB 0.0068 0.01 0.0001

170: 2,3,4,5,2',3',4'–HpCB <4.2 0.0001 0.0002

180: 2,3,4,5,2',4',5'–HpCB <4.2 0.00001 0.0000

189: 2,3,4,5,3',4',5'–HpCB 0.28 0.0001 0.0000

Others 18.85 0. 0.0000

Sum 111. g/kg 0.0264 g/kg

From Ahlborg et al. (1994)

TEQ = Conc. × TEF

Not detected, treated as half the detection limit of 4.2 g/kg

Nondioxin-like PCB congeners

The lifetime average daily dose would be recalculated separately for the dioxin-

like and nondioxin-like portions of the mixture (table 5-5):

Table 5–5. Sample lifetime average daily dose calculations (example 3)

Pathway

C IR ED BW LT LADD

Vapor inhalation 0.01 g/m 20 m /d 30 yr 70 kg 70 yr 1.2×10 mg/kg-d

33 –6

Drinking water 5. g/L 2 L/d 30 yr 70 kg 70 yr 6.1×10 mg/kg-d

–5

Fish ingestion

Dioxin TEQ 0.0264 g/kg 30 g/d 30 yr 70 kg 70 yr 4.8×10 mg/kg-d

–9

Nondioxin-like 19 g/kg 30 g/d 30 yr 70 kg 70 yr 3.5×10 mg/kg-d

–6

LADD

/ (

)

Using 150,000. per mg/kg-d as the slope for dioxin, the cancer risk would be

calculated as follows (table 5-6):

Table 5–6. Sample risk calculations (example 3)

Pathway

LADD

Slope Risk

Vapor inhalation 1.2×10 mg/kg-d 0.4 per mg/kg-d 4.8×10

–6 –7

Drinking water 6.1×10 mg/kg-d 0.07 per mg/kg-d 4.3×10

–5 –6

Fish ingestion

Dioxin TEQ 4.8×10 mg/kg-d 150,000 per mg/kg-d 7.2×10

–9 –4

Nondioxin-like PCBs 3.5×10 mg/kg-d 2 per mg/kg-d 7.0×10

–6 –6

Sum 7.3×10

–4

Risk =

LADD

× Slope

This example, although perhaps extreme, shows how it is possible for a total-

PCB approach to underestimate the toxicity of a mixture when concentrations of a few

dioxin-like or highly toxic congeners are enhanced through environmental and

metabolic processes. This shows the importance of obtaining congener analyses and

of continuing to develop quantitative methods for incorporating congener information

into risk assessments.

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APPENDIX: EMPIRICAL MODELING RESULTS

Table A–1. Empirical modeling of liver tumors in female Sprague-Dawley rats fed Aroclor 1260

Tumors Liver hepatocellular adenomas and carcinomas

Animal Female Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.403 0.413 0.399 0.391 kg

Administered dose 0 25 50 100 ppm diet

Equivalent human dose 0 0.35 0.72 1.52 mg/kg-d

Tumor incidence 1/85 10/49 11/45 24/50

Model Risk(

) = 1 – exp(–0.013–0.44

) in experimental range

Potency, slope estimates ED10=0.24, LED10=0.19, ED01=0.023, LED01=0.018,

*=0.56

Table A–2. Empirical modeling of liver tumors in female Sprague-Dawley rats fed Aroclor 1254

Tumors Liver hepatocellular adenomas and carcinomas

Animal Female Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.403 0.375 0.355 0.321 kg

Administered dose 0 25 50 100 ppm diet

Equivalent human dose 0 0.36 0.76 1.59 mg/kg-d

Tumor incidence 1/85 19/45 28/49 28/49

Model Risk(

) = 1 – exp(–0.012–1.2

) in experimental range

Potency, slope estimates ED10=0.086, LED10=0.067, ED01=0.0082, LED01=0.0064,

*=1.6

Table A–3. Empirical modeling of liver tumors in female Sprague-Dawley rats fed Aroclor 1242

Tumors Liver hepatocellular adenomas and carcinomas

Animal Female Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.403 0.398 0.372 kg

Administered dose 0 50 100 ppm diet

Equivalent human dose 0 0.75 1.53 mg/kg-d

Tumor incidence 1/85 11/49 15/45

Model Risk(

) = 1 – exp(–0.012–0.28

) in experimental range

Potency, slope estimates ED10=0.38, LED10=0.27, ED01=0.036, LED01=0.026,

*=0.39

Table A–4. Empirical modeling of liver tumors in female Sprague-Dawley rats fed Aroclor 1016

Tumors Liver hepatocellular adenomas and carcinomas

Animal Female Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.403 0.414 0.421 0.394 kg

Administered dose 0 50 100 200 ppm diet

Equivalent human dose 0 0.72 1.43 2.99 mg/kg-d

Tumor incidence 1/85 1/48 7/45 6/50

Model Risk(

) = 1 – exp(–0.012–0.044

) in experimental range

Potency, slope estimates ED10=2.4, LED10=1.4, ED01=0.23, LED01=0.14

*=0.073

Table A–5. Empirical modeling of liver tumors in male Sprague-Dawley rats fed Aroclor 1260

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.697 0.690 0.703 0.695 kg

Administered dose 0 25 50 100 ppm diet

Equivalent human dose 0 0.31 0.62 1.25 mg/kg-d

Tumor incidence 7/98 3/50 6/49 10/49

Model Risk(

) = 1 – exp(–0.071–0.11

) in experimental range

Potency, slope estimates ED10=1.0, LED10=0.55, ED01=0.31, LED01=0.053

*=0.19

Table A–6. Empirical modeling of liver tumors in male Sprague-Dawley rats fed Aroclor 1254

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.697 0.694 0.659 0.601 kg

Administered dose 0 25 50 100 ppm diet

Equivalent human dose 0 0.31 0.62 1.29 mg/kg-d

Tumor incidence 7/98 4/48 4/49 6/47

Model Risk(

) = 1 – exp(–0.075–0.0052

–0.032

) in experimental range

Potency, slope estimates ED10=1.7, LED10=0.87, ED01=0.49, LED01=0.083

*=0.12

Table A–7. Empirical modeling of liver tumors in male Sprague-Dawley rats fed Aroclor 1242

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.697 0.711 0.681 kg

Administered dose 0 50 100 ppm diet

Equivalent human dose 0 0.60 1.25 mg/kg-d

Tumor incidence 7/98 1/50 5/46

Model Risk(

) = 1 – exp(–0.058–0.012

) in experimental range

Potency, slope estimates ED10=2.9, LED10=1.2, ED01=0.91, LED01=0.16

*=0.061

Table A–8. Empirical modeling of liver tumors in male Sprague-Dawley rats fed Aroclor 1016

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Sprague-Dawley rats

Reference Brunner et al. (1996), doses from Keenan and Stickney (1996)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.697 0.723 0.699 0.712 kg

Administered dose 0 50 100 200 ppm diet

Equivalent human dose 0 0.61 1.23 2.44 mg/kg-d

Tumor incidence 7/98 2/48 2/50 4/49

Model Risk(

) = 1 – exp(–0.058–0.0031

) in experimental range

Potency, slope estimates ED10=5.9, LED10=2.5, ED01=1.8, LED01=0.32

*=0.032

Table A–9. Empirical modeling of liver tumors in female Sherman rats fed Aroclor 1260

Tumors Liver hepatocellular adenomas and carcinomas

Animal Female Sherman rats

Reference Kimbrough et al. (1975), reevaluated by Moore et al. (1994)

Exposure duration 21 mo

Study duration 23 mo (assumed animal lifespan)

Animal weight 0.35 kg (assumed)

Administered dose 0 100 ppm diet

Equivalent human dose 0 1.3 mg/kg-d

Tumor incidence 1/187 138/189

Model Risk(

) = 1 – exp(–0.0054–1.0

) in experimental range

Potency, slope estimates ED10=0.10, LED10=0.091, ED01=0.010, LED01=0.0086

*=1.2

Table A–10. Empirical modeling of liver tumors in male Fischer 344 rats fed Aroclor 1254

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Fischer 344 rats

Reference NCI (1978), reevaluated by Moore et al. (1994)

Exposure duration 104–105 wk

Study duration 104–105 wk (assumed animal lifespan)

Animal weight 0.3 kg

Administered dose 0 25 50 100 ppm diet

Equivalent human dose 0 0.32 0.64 1.28 mg/kg-d

Tumor incidence 0/24 1/24 1/24 3/23

Model Risk(

) = 1 – exp(–0.091

–0.0099

) in experimental range

Potency, slope estimates ED10=1.0, LED10=0.55, ED01=0.11, LED01=0.052,

*=0.19

Table A–11. Empirical modeling of liver tumors in female Fischer 344 rats fed Aroclor 1254

Tumors Liver hepatocellular adenomas and carcinomas

Animal Female Fischer 344 rats

Reference NCI (1978), reevaluated by Moore et al. (1994)

Exposure duration 104–105 wk

Study duration 104–105 wk (assumed animal lifespan)

Animal weight 0.2 kg

Administered dose 0 25 50 100 ppm diet

Equivalent human dose 0 0.29 0.58 1.16 mg/kg-d

Tumor incidence 0/23 1/24 2/24 1/24

Model Risk(

) = 1 – exp(–0.084

) in experimental range

Potency, slope estimates ED10=1.2, LED10=0.61, ED01=0.12, LED01=0.058,

*=0.17

Table A–12. Empirical modeling of liver tumors in male Wistar rats fed Clophen A 30

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Wistar rats

Reference Schaeffer et al. (1984), reevaluated by Moore et al. (1994)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.35 kg (assumed)

Administered dose 0 100 ppm diet

Equivalent human dose 0 1.3 mg/kg-d

Tumor incidence 8/120 16/128

Model Risk(

) = 1 – exp(–0.069–0.050

) in experimental range

Potency, slope estimates ED10=2.1, LED10=1.0, ED01=0.20, LED01=0.096,

*=0.10

Table A–13. Empirical modeling of liver tumors in male Wistar rats fed Clophen A 60

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Wistar rats

Reference Schaeffer et al. (1984), reevaluated by Moore et al. (1994)

Exposure duration 24 mo

Study duration 24 mo (assumed animal lifespan)

Animal weight 0.35 kg (assumed)

Administered dose 0 100 ppm diet

Equivalent human dose 0 1.3 mg/kg-d

Tumor incidence 8/120 114/125

Model Risk(

) = 1 – exp(–0.069–1.8

) in experimental range

Potency, slope estimates ED10=0.058, LED10=0.047, ED01=0.0055, LED01=0.0045,

*=2.2

Table A–14. Empirical modeling of liver tumors in male Sprague-Dawley rats fed Aroclor 1260

Tumors Liver hepatocellular adenomas and carcinomas

Animal Male Sprague-Dawley rats

Reference Norback and Weltman (1985), reevaluated by Moore et al. (1994)

Exposure duration 24 mo

Study duration 29 mo (assumed animal lifespan)

Animal weight 0.35 kg (assumed)

Administered dose 0 100 ppm diet

Equivalent human dose 0 1.3 mg/kg-d

Tumor incidence 0/31 5/40

Model Risk(

) = 1 – exp(–0.10

) in experimental range

Potency, slope estimates ED10=1.0, LED10=0.53, ED01=0.098, LED01=0.051,

*=0.20

Table A–15. Empirical modeling of liver tumors in female Sprague-Dawley rats fed Aroclor 1260

Tumors Liver hepatocellular adenomas and carcinomas

Animal Female Sprague-Dawley rats

Reference Norback and Weltman (1985), reevaluated by Moore et al. (1994)

Exposure duration 24 mo

Study duration 29 mo (assumed animal lifespan)

Animal weight 0.35 kg (assumed)

Administered dose 0 100 ppm diet

Equivalent human dose 0 1.3 mg/kg-d

Tumor incidence 1/45 41/46

Model Risk(

) = 1 – exp(–0.022–1.7

) in experimental range

Potency, slope estimates ED10=0.062, LED10=0.046, ED01=0.0059, LED01=0.0044,

*=2.3