SELECTED PCBs
(AROCLOR -1260, -1254,
-1248, -1242, -1232, -1221,
and-1016)
.'-9
*
Agency for Toxic Substances and Disease Registry
U.S. Public Health Service
-------
ATSDR/TP-88/21
TOXICOLOGICAL PROFILE FOR
SELECTED PCBs
(Aroclor-1260, -1254, -1248, -1242,
-1232, -1221, and -1016)
Date Published June 1989
Prepared by:
Syracuse Research Corporation
under Contract No. 68-C8-0004
for
Agency for Toxic Substances and Disease Registry (ATSDR)
U.S. Public Health Service
in collaboration with
U.S. Environmental Protection Agency (EPA)
Technical editing/document preparation by:
Oak Ridge National Laboratory
under
DOE Interagency Agreement No. 1857-B026-A1
-------
DISCLAIMER
Mention of company name or product does not constitute endorsement by
the Agency for Toxic Substances and Disease Registry.
-------
FOREWORD
The Superfund Amendments and Reauthorization Act of 1986 (Public
Law 99-499) extended and amended the Comprehensive Environmental
Response, Compensation, and Liability Act of 1980 (CERCLA or Superfund).
This public law (also known as SARA) directed the Agency for Toxic
Substances and Disease Registry (ATSDR) to prepare toxicological
profiles for hazardous substances which are most commonly found at
facilities on the CERCLA National Priorities List and which pose the
most significant potential threat to human health, as determined by
ATSDR and the Environmental Protection Agency (EPA). The list of the 100
most significant hazardous substances was published in the Federal
Register on April 17, 1987.
Section 110 (3) of SARA directs the Administrator of ATSDR to
prepare a toxicological profile for each substance on the list. Each
profile must include the following content:
"(A) An examination, summary, and interpretation of available
toxicological information and epidemiologic evaluations on a
hazardous substance in order to ascertain the levels of significant
human exposure for the substance and the associated acute,
subacute, and chronic health effects.
(B) A determination of whether adequate information on the health
effects of each substance is available or in the process of
development to determine levels of exposure which present a
significant risk to human health of acute, subacute, and chronic
health effects.
(C) Where appropriate, an identification of toxicological testing
needed to identify the types or levels of exposure that may present
significant risk of adverse health effects in humans."
This toxicological profile is prepared in accordance with
guidelines developed by ATSDR and EPA. The guidelines were published in
the Federal Register on April 17, 1987. Each profile will be revised and
republished as necessary, but no less often than every three years, as
required by SARA.
The ATSDR toxicological profile is intended to characterize
succinctly the toxicological and health effects information for the
hazardous substance being described. Each profile identifies and reviews
the key literature that describes a hazardous substance's toxicological
properties. Other literature is presented but described in less detail
than the key studies. The profile is not intended to be an exhaustive
document; however, more comprehensive sources of specialty information
are referenced.
iii
-------
Foreword
Each toxlcological profile begins with a public health statement,
which describes in nontechnical language a substance's relevant
toxicological properties. Following the statement is material that
presents levels of significant human exposure and, where known,
significant health effects. The adequacy of information to determine a
substance's health effects is described in a health effects summary.
Research gaps in toxicologic and health effects information are
described in the profile. Research gaps that are of significance to
protection of public health will be identified by ATSDR, the National
Toxicology Program of the Public Health Service, and EPA. The focus of
the profiles is on health and toxicological information; therefore, we
have included this information in the front of the document.
The principal audiences for the toxicological profiles are health
professionals at the federal, state, and local levels, interested
private sector organizations and groups, and members of the public. We
plan to revise these documents in response to public comments and as
additional data become available; therefore, we encourage comment that
will make the toxicological profile series of the greatest use.
This profile reflects our assessment of all relevant toxicological
testing and information that has been peer reviewed. It has been
reviewed by scientists from ATSDR, EPA, the Centers for Disease Control,
and the National Toxicology Program. It has also been reviewed by a
panel of nongovernment peer reviewers and was made available for public
review. Final responsibility for the contents and views expressed in
this toxicological profile resides with ATSDR.
James 0. Mason, M.D., Dr. P.M.
Assistant Surgeon General
Administrator, ATSDR
iv
-------
CONTENTS
FOREWORD
LIST OF FIGURES ix
LIST OF TABLES xi
1. PUBLIC HEALTH STATEMENT 1
1.1 WHAT ARE PCBs? 1
1.2 HOW MIGHT I BE EXPOSED TO PCBs? 1
1.3 HOW DO PCBs GET INTO MY BODY? 2
1.4 HOW DO PCBs AFFECT MY HEALTH? 2
1.5 IS THERE A MEDICAL TEST TO DETERMINE IF I HAVE BEEN
EXPOSED TO PCBs? 3
1.6 WHAT LEVELS OF EXPOSURE HAVE RESULTED IN HARMFUL
HEALTH EFFECTS? 3
1.7 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT
MADE TO PROTECT HUMAN HEALTH? 3
2. HEALTH EFFECTS SUMMARY 9
2.1 INTRODUCTION 9
2.2 LEVELS OF SIGNIFICANT EXPOSURE 10
2.2.1 Key Studies and Graphical Presentations 10
2.2.1.1 Inhalation 17
2.2.1.2 Oral 18
2.2.1.3 Dermal 21
2.2.2 Biological Monitoring as a Measure of
Exposure and Effects 22
2.2.2.1 Exposure 22
2.2.2.2 Effects 25
2.2.3 Environmental Levels as Indicators of
Exposure and Effects 30
2.2.3.1 Levels found in the environment 30
2.2.3.2 Human exposure potential 30
2.3 ADEQUACY OF DATABASE 31
2.3.1 Introduction 31
2.3.2 Health Effect End Points 32
2.3.2.1 Introduction and graphic summary 32
2.3.2.2 Descriptions of highlights of graphs 32
2.3.2.3 Summary of relevant ongoing research .... 35
2.3.3 Other Information Needed for Human
Health Assessment 35
2.3.3.1 Pharmocokinetics and mechanisms of
action 35
2.3.3.2 Monitoring of human biological samples .. 35
2.3.3.3 Environmental considerations 35
-------
Concents
CHEMICAL AND PHYSICAL INFORMATION 37
3.1 CHEMICAL IDENTITY " ' 37
3.2 PHYSICAL AND CHEMICAL PROPERTIES " ' ' ' 37
TOXICOLOGICAL DATA 43
4.1 OVERVIEW 43
4.2 TOXICOKINETICS '.'.'.'.'.'.'. 45
4.2.1 Absorption 45
4.2.1.1 Inhalation 45
4.2.1.2 Oral., 45
4.2.1.3 Dermal 46
4.2.2 Distribution 46
4.2.2.1 Inhalation 46
4.2.2.2 Oral 47
4.2.2.3 Dermal 49
4.2.3 Metabolism 49
4.2.3.1 Human 49
4.2.3.2 Animal 49
4.2.4 Excretion 50
4.2.4.1 Inhalation 50
4.2.4.2 Oral 50
4.2.4.3 Dermal 52
4.2.4.4 Parenteral routes 52
4.3 TOXICITY 52
4.3.1 Lethality and Decreased Longevity 52
4.3.1.1 Inhalation 52
4.3.1.2 Oral 53
4.3.1.3 Dermal 55
4.3.2 Systemic/Target Organ Toxicity 55
4.3.2.1 Liver 55
4.3.2.2 Cutaneous tissues 60
4.3.2.3 Immunological effects 62
4.3.2.4 Thyroid 64
4.3.2.5 Stomach 65
4.3.2.6 Porphyria 65
4.3.2.7 Kidney 67
4.3.3 Developmental Toxicity 67
4.3.3.1 Inhalation 67
4.3.3.2 Oral 67
4.3.3.3 Dermal 70
4.3.3.4 General discussion 70
4.3.4 Reproductive Toxicity 70
4.3.5 Genotoxicity 71
4.3.5.1 Human 71
4.3.5.2 Animal 71
4.3.6 Carcinogenicity 73
4.3.6.1 Inhalation 73
4.3.6.2 Oral 74
4.3.6.3 Dermal 77
4.3.6.4 General discussion 77
4.4 INTERACTIONS WITH OTHER CHEMICALS 79
vi
-------
Concents
5. MANUFACTURE, IMPORT. USE, AND DISPOSAL 81
5.1 OVERVIEW 81
5.2 PRODUCTION 81
5.3 IMPORT 81
5.4 USES 82
5.5 DISPOSAL '.'.'.'.'.'.'.'.'.'.'.'.'.'.'.'. 82
6. ENVIRONMENTAL FATE 83
6.1 OVERVIEW '.'.'.'.'.'.'.'.'.'.'. 83
6.2 RELEASES TO THE ENVIRONMENT '.'.'.'. 83
6.3 ENVIRONMENTAL FATE 84
6.3.1 Transport and Partitioning 84
6.3.2 Transformation and Degradation 85
7. POTENTIAL FOR HUMAN EXPOSURE 87
7.1 OVERVIEW 87
7.2 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT 87
7.2.1 Air 87
7.2.2 Water 88
7.2.3 Soil 89
7.2.4 Other 89
7.2.4.1 Foodstuffs 89
7.2.4.2 Fish and precipitation 91
7.3 OCCUPATIONAL EXPOSURES 93
7.4 POPULATIONS AT HIGH RISK 93
8. ANALYTICAL METHODS 95
8.1 ENVIRONMENTAL MEDIA 95
8.2 BIOMEDICAL SAMPLES 95
9. REGULATORY AND ADVISORY STATUS 99
9.1 INTERNATIONAL 99
9.2 NATIONAL 99
9.2.1 Regulations 99
9.2.1.1 Air 99
9.2.1.2 Food 99
9.2.1.3 Water 99
9.2.2 Advisory Guidance 99
9.2.2.1 Air 99
9.2.2.2 Water 100
9.2.2.3 Soil 100
9.2.2.4 Others 100
9.2.3 Data Analysis 100
9.3 STATE 101
10. REFERENCES 103
11. GLOSSARY 131
APPENDIX: PEER REVIEW 135
vil
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LIST OF FIGURES
1.1 Health effects from breathing PCBs 4
1.2 Health effects from ingesting PCBs 5
1.3 Health effects from skin contact with PCBs 6
2.1 Effects of PCBs-- inhalation exposure 11
2.2 Effects of PCBs--oral exposure 12
2.3 Effects of PCBs--dermal exposure 13
2.4 Levels of significant exposure for PCBs--inhalation 14
2.5 Levels of significant exposure for PCBs--oral IS
2.6 Levels of significant exposure for PCBs--dermal 16
2.7 Availability of information on health effects of PCBs
(human data) 33
2.8 Availability of information on health effects of PCBs
(animal data) 34
ix
-------
LIST OF TABLES
2.1 PCB levels in blood of exposed workers
(Aroclors 1016 , 1242 , 1248) .................................. 24
2.2 PCB blood levels (Aroclor 1254) and duration of exposure ..... 24
2.3 Serum PCB concentrations in U.S. populations without
occupational exposure to PCBs and in subpopulations
consuming fish from PCB -contaminated waters .................. 26
2.4 Serum PCB concentrations in populations with occupational
exposure [[[ 28
3 . 1 Chemical identity of the Aroclors ............................ 39
3 . 2 Physical and chemical properties of PCBs ..................... 40
3 . 3 Approximate molecular composition of PCBs .................... 42
4 . 1 Acute oral LDcnS of Aroclors ................................. 54
4.2 Acute dermal LD50 values of Aroclors in rabbits .............. 56
4 . 3 Genotoxicity of PCBs in vitro ................................ 72
4 . 4 Genotoxicity of PCBs in vivo ................................. 72
7.1 Aroclor residues in raw domestic agricultural commodities
for fiscal years 1970-1976 ................................... 90
7.2 Estimated dietary intake of PCBs for adults, infants, and
toddlers (pAg/day) ......................................... 92
8 . 1 Analytical methods for environmental media ................... 96
8 . 2 Analytical methods for biological samples .................... 97
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1. PUBLIC HEALTH STATEMENT
1.1 WHAT ARE PCBs?
The abbreviation PCB refers to polychlorinated biphenyls. PCBs are
a family of man-made chemicals that contain 209 individual compounds
with varying toxicity. Commercial formulations of PCBs enter the
environment as mixtures consisting of a variety of PCBs and impurities.
Because of the complex nature associated with evaluating the health
effects of PCBs, this document will address only seven selected classes
of PCBs, which include 35% of all of the different PCBs and 98% of PCBs
sold in the United States since 1970. Some commercial PCB mixtures are
known in the United States by their industrial trade name, Aroclor.
Because of their insulating and nonflammable properties, PCBs have been
used widely as coolants and lubricants in transformers, capacitors, and
other electrical equipment. The manufacture of PCBs stopped in the
United States in October 1977 because of evidence that PCBs accumulate
in the environment and may cause health hazards for humans.
1.2 HOV NIGHT I BE EXPOSED TO PCBs?
Although PCBs are no longer manufactured, human exposure still
occurs. Many older transformers and capacitors still contain fluids that
contain PCBs. The useful lifetime of many of these transformers can be
30 years or more.
The two main sources of human exposure to PCBs are environmental
and occupational. PCBs are very persistent chemicals that are widely
distributed throughout the entire environment. PCBs have been found in
at least 216 of 1,177 hazardous waste sites on the National Priorities
List (NPL). Background levels of PCBs can be found in the outdoor air,
on soil surfaces, and in water. Eating contaminated fish can be a major
source of PCB exposure to humans. These PCBs originate in contaminated
water, sediment, PCB-laden particulates, and in fish that have eaten
PCB-contaminated prey. Although PCBs found in fish are generally
concentrated in nonedible portions, the amounts in edible portions are
high enough to make consumption a major source of exposure for humans.
Compared with the intake of PCBs through eating contaminated fish,
exposure through breathing outdoor air containing PCBs is small. Most of
the PCBs in outdoor air may be present because of an environmental
cycling process. PCBs in water, or on soil surfaces, evaporate and are
then returned to earth by rainfall or settling of dust particles.
Reevaporation repeats the cycle. Once In the air, PCBs can be carried
long distances; they have been found in snow and seawater in the
Antarctic. In addition, contaminated indoor air may be a major source of
human exposure to PCBs, particularly In buildings that contain PCB-
containing devices.
-------
2 Section 1
PCBs can be released into the environment from:
poorly maintained toxic waste sites that contain PCBs,
illegal or improper dumping of PCB wastes, such as transformer
fluids,
leaks or fugitive emissions from electrical transformers containing
PCBs, and
disposal of PCB-containing consumer products into municipal
landfills rather than into landfills designed to hold hazardous
wastes.
Consumer products that may contain PCBs are:
old fluorescent lighting fixtures and
electrical devices or appliances containing PCB capacitors made
before PCB use was stopped.
Occupational exposure to PCBs can occur during:
repair or maintenance of PCB transformers,
accidents or spills involving PCB transformers,
disposal of PCB materials, and
contact at hazardous waste sites.
1.3 HOY DO PCBs GET INTO MY BODY?
PCBs enter the body through contaminated food and air and through
skin contact. The most common route of exposure is by eating fish and
shellfish from PCB-contaminated water. Exposure from drinking water is
minimal. It is known that nearly everyone has PCBs in their bodies,
including infants who drink breast milk containing PCBs.
1.4 HOW DO PCBs AFFECT MY HEALTH?
Although PCBs have not been manufactured in the U.S. since October
1977, their diminishing but continued presence in certain commercial
applications and trade have resulted in low-level exposure to the
general population. Prior to 1977, certain occupational settings had,
and may still have, higher levels of human exposure. Animal experiments
have shown that some PCB mixtures produce adverse health effects that
include liver damage, skin irritations, reproductive and developmental
effects, and cancer. Therefore, it is prudent to consider that there may
be health hazards for humans. Human studies to date show that
initiations, such as acnelike lesions and rashes, can occur in PCB-
exposed workers. Other studies of people with occupational exposure
suggest that PCBs might cause liver cancer. Reproductive and
developmental effects may also be related to occupational exposure and
eating of contaminated fish, while the role of PCBs in producing cancer,
reproductive, and developmental effects in humans cannot be clearly
delineated, the suggestive evidence provides an additional basis for
public health concern about humans who may be exposed to PCBs. The
complexity of relating the specific mixtures for which data are
available to exposures in the general population has resulted in a
-------
Public Health Statement 3
tendency to regard all PCBs as having a similar health hazard potential.
although this assumption may not be true.
1.3 IS THERE A MEDICAL TEST TO DETERMINE IF I
HAVE BEEN EXPOSED TO PCBs?
There are tests to determine PCBs in the blood, body fat, and
breast milk. These tests are not routine clinical tests, but they can
detect PCBs in members of the general population as well as in workers
with occupational exposure to PCBs. Although these tests indicate if one
has been exposed to PCBs, they do not predict potential health effects.
Blood tests are the easiest, safest, and, perhaps, the best method for
detecting recent large exposures. It should be recognized that nearly
everyone has been exposed to PCBs because they are found throughout the
environment and that nearly all persons are likely to have detectable
levels of PCBs in their blood, fat. and breast milk.
1.6 WHAT LEVELS OF EXPOSURE HAVE RESULTED IN HARMFUL HEALTH EFFECTS?
Figures 1.1, 1.2, and 1.3 on the following pages show the
relationship between exposure to PCBs and known health effects for the
PCBs that are covered by this profile. Other PCBs may have different
toxic properties. In the first set of graphs, labeled "Health effects
from breathing PCBs," exposure is measured in milligrams of PCBs per
cubic meter of air (mg/m3). In the second and third sets of graphs, the
same relationship is represented for the known "Health effects from
ingesting PCBs" and "Health effects from skin contact with PCBs."
Exposures are measured in milligrams of PCBs per kilogram of body weight
per day (mgAg/day). It should be noted that health effects observed by
one route of exposure may be relevant to other routes of exposure.
In all graphs, effects in animals are shown on the left side,
effects in humans on the right. The first column on the graphs, labeled
short-term, refers to known health effects from exposure to PCBs for
2 weeks or less. The columns labeled long-term refer to PCB exposures of
longer than 2 weeks. The levels marked on the graphs as anticipated to
be associated with minimal risk of developing health effects are based
on information generated from animal studies; therefore, some
uncertainty still exists. Based on evidence that PCBs cause cancer in
animals, the Environmental Protection Agency (EPA) considers PCBs to be
probable cancer-causing chemicals in humans and has estimated that
ingestion of 1 microgram of PCB per kilogram per day for a lifetime
would result in 77 additional cases of cancer in a population of 10,000
people or equivalently. 77,000 additional cases of cancer in a
population of 10,000,000 people. These risk values are plausible upper-
limit estimates. Actual risk levels are unlikely to be higher and may be
lower. J
1.7 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE
TO PROTECT HUMAN HEALTH?
For exposure via drinking water, EPA advises that the following
concentrations of PCB 1016 are levels at which adverse health effects
would not be expected: 0.0035 milligrams PCB 1016 per liter of water for
adults and 0.001 milligrams PCB 1016 per liter of water for children
-------
Section 1
SHORT-TERM EXPOSURE
(LESS THAN OR EQUAL TO 14 DAYS)
LONG-TERM EXPOSURE
(GREATER THAN 14 DAYS)
EFFECTS CONC. IN EFFECTS EFFECTS CONG IN EFFECTS
IN AIR IN ' IN AIR IN
ANIMALS (mg/m3) HUMANS ANIMALS (mg/m3) HUMANS
QUANTITATIVE
DATA WERE
NOT AVAILABLE
QUANTITATIVE ,
DATA WERE
NOT AVAILABLE
1
1
t
i
LIVER DAMAGE
1
0
0
n
) ,
1
i
\
>
1
1
01
V»1
SKIN
IRRITATION
Fig. 1.1. Health effects from bratUaf PCBa.
-------
Public Health Statement 5
SHORT-TERM EXPOSURE
(LESS THAN OR EQUAL TO 14 DAYS)
LONG-TERM EXPOSURE
(GREATER THAN 14 DAYS)
EFFECTS EFFECTS EFFECTS
IN DOSE IN IN DOSE
ANIMALS (mg/Kg/day) HUMANS ANIMALS (mg/kg/day)
OPATH ._. 750 QUANTITATIVE LIVER AND $KIN ni
1 DATA WERE DAMAGE. DEATH
₯ NOT
1 AVAILABLE
1
1
EFFECTS ON
UNBORN
5
0
5
I
1
r °
LIVER DAMAGE 0
- 0
0
EFFECTS ON
UNBORN AND <
8 NEWBORN
6
4
2
I
O.Q04 MINIMAL RISK
'
0
0
0
0
0
0
0
0
09
08
07
06
05
04
03
02
001
FOR EFFECTS *
OTHER THAN f
CANCER
o.ooi o'oooi
EFFECTS
IN
HUMANS
QUANTITATIVE
DATA WERE
NOT
AVAILABLE
MINIMAL RISK
FOR EFFECTS
OTHER THAN
CANCER
Fig. 1.2. Health effects from iogestiog PCBs.
-------
Section 1
SHORT-TERM EXPOSURE
(LESS THAN OR EQUAL TO 14 DAYS)
LONG-TERM EXPOSURE
(GREATER THAN 14 DAYS)
EFFECTS
IN
ANIMALS
DOSE
(mg/kg/day)
1400
DEATH.
EFFECTS
IN
HUMANS
OUANTrTATIVE
DATA WERE
NOT AVAILABLE
EFFECTS
IN
ANIMALS
1200
1000
800
600
400
200
DOSE
(mg/Kq/day)
1400
1200
1000
800
600
400
200
EFFECTS
IN
HUMANS
UVER AND
KIDNEY DAMAGE.
HEALTH EFFECTS
FROM SKIN CON-
TACT INCLUDE
SKIN IRRITATION
AND UVER
EFFECTS. BUT
DOSES ARE NOT
KNOWN
Flf. 1J. Hodtk effects tnm ikfe cortMt wttfc PCB*.
-------
Public Health Statement 1
The EPA has also developed guidelines for the concentrations of PCBs in
ambient water (e.g., lakes and rivers) and in drinking water that are
associated with a risk of developing cancer. The guideline for ambient
water is a range, 0.0079 to 0.79 nanograms of PCBs per liter of water,
which reflects the increased risk of one person developing cancer in
populations of 10,000,000 to 100,000 people. The guideline for drinking
water is a range, 0.005 to 0.5 micrograms of PCBs per liter of water,
which also reflects the risk of one person developing cancer in
populations of 10,000,000 to 100.000 people.
The Food and Drug Administration (FDA) specifies PCB concentration
limits of 0.2 to 3 parts per million (milligrams PCB per kilogram of
food) in infant foods, eggs, milk (in milk fat), and poultry (fat).
The National Institute for Occupational Safety and Health (NIOSH)
recommends an occupational exposure limit for all PCBs of 0.001
milligram of PCBs per cubic.meter of air (mg/m3) for a 10-hpur workday
40-hour workweek. The Occupational Safety and Health Administration
(OSHA) permissible occupational exposure limits are 0.5 and 1 0 mg/m3
for specific PCBs for an 8-hour workday.
-------
2. HEALTH EFFECTS SUMMARY
2.1 INTRODUCTION
This section summarizes and graphs data on the health effects
concerning exposure to PCBs. The purpose of this section is to present
levels of significant exposure for PCBs based on key toxicological
studies, epidemiological investigations, and environmental exposure
data. The information presented in this section is critically evaluated
and discussed in Sect. 4, Toxicological Data, and Sect. 7, Potential for
Human Exposure.
This Health Effects Summary section comprises two major parts.
Levels of Significant Exposure (Sect. 2.2) presents brief narratives and
graphics for key studies in a manner that provides public health
officials, physicians, and other interested individuals and groups with
(1) an overall perspective of the toxicology of PCBs and (2) a
summarized depiction of significant exposure levels associated with
various adverse health effects. This section also includes information
on the levels of PCBs that have been monitored in human fluids and
tissues and information about levels of PCBs found in environmental
media and their association with human exposures.
The significance of the exposure levels shown on the graphs may
differ depending on the user's perspective. For example, physicians
concerned with the interpretation of overt clinical findings in exposed
persons or with the identification of persons with the potential to
develop such disease may be interested in levels of exposure associated
with frank effects (Frank Effect Level, PEL). Public health officials
and project managers concerned with response actions at Superfund sites
may want information on levels of exposure associated with more subtle
effects in humans or animals (Lowest-Observed-Adverse-Effect Level,
LOAEL) or exposure levels below which no adverse effects (No-Observed-
Adverse-Effect Level, NOAEL) have been observed. Estimates of levels
posing minimal risk to humans (Minimal Risk Levels, MRL) are of interest
to health professionals and citizens alike.
Adequacy of Database (Sect. 2.3) highlights the availability of key
studies on exposure to PCBs in the scientific literature and displays
these data in three-dimensional graphs consistent with the format in
Sect. 2.2. The purpose of this section is to suggest where there might
be insufficient information to establish levels of significant human
exposure. These areas will be considered by the Agency for Toxic
Substances and Disease Registry (ATSDR), EPA, and the National
Toxicology Program (NTP) of the U.S. Public Health Service in order to
develop a research agenda for PCBs.
-------
10 Section 2
2.2 LEVELS OF SIGNIFICANT EXPOSURE
To help public health professionals address the needs of persons
living or working near hazardous waste sites, the toxicology data
summarized in this section are organized first by route of exposure--
inhalation, ingestion, and dermal--and then by toxicological end points
that are categorized into six general areas--lethality, systemic/target
organ toxicity, developmental toxicity, reproductive toxicity, genetic
toxicity, and careinogenieity. The data are discussed in terms of three
exposure periods--acute, intermediate, and chronic.
Two kinds of graphs are used to depict the data. The first type is
a "thermometer" graph. It provides a graphical summary of the human and
animal toxicological end points (and levels of exposure) for each
exposure route for which data are available. The ordering of effects
does not reflect the exposure duration or species of animal tested. Tho
second kind of graph shows Levels of Significant Exposure (LSE) for each
route and exposure duration. The points on the graph showing NOAELs and
LOAELs reflect the actual doses (levels of exposure) used in the key
studies. No adjustments for exposure duration or intermittent exposure
protocol were made.
Adjustments reflecting the uncertainty of extrapolating animal data
to man, intraspecies variations, and differences between experimental vs
actual human exposure conditions were considered when estimates of
levels posing minimal risk to human health were made for noncancer end
points. These minimal risk levels were derived for the most sensitive
noncancer end point for each exposure duration by applying uncertainty
factors. These levels are shown on the graphs as a broken line starting
from the actual dose (level of exposure) and ending with a concave-
curved line at its terminus. Although methods have been established to
derive these minimal risk levels (Barnes et al. 1987), shortcomings
exist in the techniques that reduce the confidence in the projected
estimates. Also shown on the graphs under the cancer end point are low-
level risks (10~4 to 10*7) reported by EPA. In addition, the actual dose
(level of exposure) associated with the tumor incidence is plotted.
Evaluation of the toxicity of Aroclors and other commercial PCB
mixtures is complicated by numerous considerations. Because of these
considerations, it is assumed, for the purpose of health effects
evaluation, that effects resulting from exposure to a specific Aroclor
are representative of effects that may be produced by the other Aroclors
(see discussion in preface to Sect. 4.3).
2.2.1 Key Studies and Graphical Presentations
Dose*response-duration data for the toxicity and carcinogenicity of
the PCBs discussed in this profile are displayed in two types of graphs.
These data are derived from the key studies described in the following
sections. The "thermometer" graphs in Figs. 2.1, 2.2, and 2.3 plot
exposure levels vs NOAELs and LOAELs for various effects and durations
of inhalation, oral, and dermal exposures, respectively. The graphs of
levels of significant exposure in Figs. 2.4, 2.S, and 2.6 plot end-
point-specific NOAELs. LOAELs, and/or minimal levels of risk for acute
(-514 days), intermediate (15-364 days), and chronic (2365 days)
durations for inhalation, oral, and dermal exposures, respectively.
-------
Health Effects Suonary 11
ANIMALS
10
O RAT MOUSE. RABBIT GUINEA PIG. CAT.
NO DEATHS. 24 DAYS. INTERMTTENT
RAT MOUSE. RABBIT. GUINEA PIG CAT.
LJVEH TOXICITY. 213 DAYS INTERMITTENT
01
001
0001 I-
HUMANS
(mgftn1)
10 .-
01
001
0001
SKIN AND
POSSIBLY
UVER EFFECTS
LOAEL
O NQAEL
Fig. 2.1. Effects of PCBs lahdatioa
-------
12 Section 2
ANIMALS
HUMANS
1000
100
10
RAT LDso. SINGLE DOSE
MNK. LOso. SINGLE DOSE
O MOUSE. DEVELOPMENTAL TOX1CITY. 1DOSE
MOUSE. DEATH. 14 DAYS. CONTINUOUS
O MOUSE. DEATH. 14 DAYS. CONTINUOUS
RABBIT. DEVELOPMENTAL TOXICITY. 28 DAYS. CONTINUOUS
_ O RABBIT. DEVELOPMENTAL TOXICITY. 28 DAYS, CONTINUOUS
MINK. DIETARY LDso. 28 DAYS. CONTINUOUS
RAT. DEVELOPMENTAL TOXICITY. 21 DAYS. CONTINUOUS
MINK. DIETARY LDso. MONTHS. CONTINUOUS: RAT. DECREASED LONGEVITY.
104 WEEKS. CONTINUOUS
RAT. REPRODUCTIVE TOXICITY. 1-2 GENERATIONS. CONTINUOUS
RAT. UVER TOXICITY. 4 DAYS. CONTINUOUS
1
O RAT. UVER TOXICITY. 4 DAYS. CONTINUOUS
MINK, REPRODUCTIVE TOXICITY. 170 DAYS. CONTINUOUS
RAT. UVER TOXICITY. 2-6 MONTHS. CONTINUOUS
O RAT. REPRODUCTIVE TOXICITY. 1-2 GENERATIONS. CONTINUOUS
0 i |_ MONKEY. UVER AND SKIN TOXICITY. 173 DAYS. CONTINUOUS
MONKEY. DEVELOPMENTAL TOXICITY. 87 WEEKS. CONTINUOUS
O RAT. UVER TOXICITY. 4 WEEKS, CONTINUOUS
001 I O MONKEY. DEVELOPMENTAL TOXICITY. 87 WEEKS CONTINUOUS
LOAEL ONOAEL
QUANTITATIVE DATA
WERE NOT AVAILABLE
Fig. 2.2. Effects of PCBeonl exponre.
-------
Health Effects Summary 13
ANIMALS
(mg/kg/day)
HUMANS
10.000 i
1.000
100
10
RABBIT. LD-. SINGLE DOSE
RABBIT. LJVER. KIDNEY. AND SKIN TOXICITY. 38 DAYS. INTERMITTENT
OCCUPATIONAL
EXPOSURE IS
ASSOCIATED WITH
LIVER AND SKIN
EFFECTS. BUT
DERMAL DOSES ARE
NOT AVAILABLE
LOAEL
Fig. 2J. Effects of PCBsdermal exposure.
-------
14 Section 2
ACUTE
(£ 14 DAYS)
INTERMEDIATE
(15-364 DAYS)
CHRONIC
365 DAYS)
(mg/m3)
DECREASED
LONGEVITY
10
QUANTITATIVE
DATA WERE NOT OQ.h.m.r.c
AVAILABLE (AROCLOR1242)
TARGET
ORGAN
O g, h. m, r. c
(UVER)
(AROCLOR 1242)
TARGET
ORGAN
1
Q, h, m. r, c (LIVER)
(AROCLOR 1254)
0.1
0.01
(SKIN AND
POSSIBLY LIVER)
(AROCLORS 1242
AND 1254)
1
0.001 L-
LOAEL IN ANIMALS
O NOAEL IN ANIMALS
1 RANGE OF EFFECT
FOR HUMANS
g GUINEA PIG
h RABBIT
m MOUSE
r RAT
c CAT
for
-------
Health Effects Siunary IS
ACUTE
(S 14 DAYS)
LETHALITY
DEVELOP
MENTAL
TARGET
ORGAN
DECREASED
LONGEVITY
INTERMEDIATE
(15-364 DAYS)
TARGET
ORGAN
REPRO-
DUCTION
DEVELOP-
MENTAL
CHRONIC
(2365 DAYS)
DECREASED
LONGEVITY
CANCER
(mg/kg/day)
1000
100
10
01
001
0001
00001
000001
0000001
0 0000001
0 00000001
_ «r (AROCLOR 1254)
n (AROCLOR 1221)
OTHAROCLOR
1254)
- in (AROCLOR 1254)
r
|h(AROCLOR12S4)
»r(AROCLOR1254)
n (AROCLOR
r (LIVER) 1254)
I (AROCLOR
I 1254)
ir (LIVER)
I
r (AROCLOR 1254)
n(AROCLOR1254)
r (AROCLOR
1254)
»r (AROCLOR
1260)
k (LIVER)
(AROCLOR 1248)
6(AROCLORS 1242
1248 1254 1260)
k (AROCLOR 1016)
S>
10-*-i
ID'1
5_
MINIMAL RISK LEVEL
FOR EFFECTS OTHER
THAN CANCER
I MINIMAL RISK LEVEL
J EXTRAPOLATED FROM
INTERMEDIATE
EXPOSURE DATA
r RAT
n MINK
m MOUSE
k MONKEY
n RABBfT
LOAEL
ONOAEL
10-
LOAEL AND
I NOAEL IN
U SAME SPECIES
7_
ESTIMATED
UPPER-
BOUND
HUMAN
CANCER
RISK LEVELS
Flf.24. L«t«toofrig«iflc»«t
for
-------
16 Section 2
ACUTE INTERMEDIATE CHRONIC
(S 14 DAYS) (15-364 DAYS) & 365 DAYS)
LETHALITY TARGET ORGAN
(mg/kg/day)
10.000 |- QUANTITATIVE DATA
WERE NOT AVAILABLE
1.000
100
10 «-
h (AROCLOR
1221)
h (LIVER. KIDNEY,
SKIN) (AROCLOR
1260)
LOAEL
h RABBIT
Fig. 2.6. Lercb of «ifMt IKBOMM for
-------
Health Effects Summary 17
Dermal exposure contributes significantly to occupational exposure, but
the relative contributions of dermal and inhalation exposure in
occupational settings has not been discerned (Wolff 1985). Furthermore
occupational exposure levels are expressed as concentrations of PCBs in
air, making it difficult to quantitate dermal exposure doses. For this
reason, effects of occupational exposure are discussed under inhalation
exposure and plotted in Figs. 2.1 and 2.4 (graphs for inhalation
exposure).
2.2.1.1 Inhalation
Lethality and decreased longevity. Data regarding inhalation
exposure levels that produce death in humans were not available.
Exposure to near saturation vapor concentrations of heated Aroclor 1242
(8.6 mg/m->) 7 h/day, 5 days/week for 24 days was not lethal for cats,
rats, mice, rabbits, or guinea pigs (Treon et al. 1956). This
concentration represents a NOAEL for lethality for intermediate
inhalation exposures (see Figs. 2.1 and 2.4). No data were available
regarding lethality/decreased longevity of animals due to acute or
chronic inhalation exposure to PCBs.
Systemic/target organ toxicity. Oral toxicity studies in animals
have established that the liver and cutaneous tissues are primary target
organs of PCBs. Human health surveys have associated occupational
exposure to PCBs with increased serum levels of liver-associated enzymes
and dermatologic effects such as chloracne and skin rashes (Sects.
4.3.2.1 and 4.3.2.2). The results of some of these studies are
equivocal, and exposure levels were not reported or inadequately
characterized. Furthermore, although inhalation is considered a major
route of exposure, the contribution of dermal exposure to total
occupational exposure is also significant.
Fischbein et al. (1979, 1982, 1985) reported data suggestive of
associations between serum levels of PCBs and SCOT levels and
dermatologic effects in workers who had been exposed to 8-h time-
weighted average concentrations of Aroclors, primarily 1242 and 1254,
ranging from 0.007-11.0 mg/m3. Because of limitations of this study
(Sects. 4.3.2.1 and 4.3.2.2), these effects could be regarded as
inconclusive and cannot be associated with specific exposure
concentrations. It is, however, appropriate to plot the range of Aroclor
concentrations from this study in Figs. 2.1 and 2.4 because similar
effects have been observed in other health surveys of PCB-exposed
workers, information regarding human liver histopathology is lacking,
and the liver and skin are unequivocal targets of PCB toxicity in
animals. This concentration range is intended to approximate typical
concentrations in occupational environments that may be associated with
hepatic and dermatologic alterations.
In the only animal inhalation study of PCBs, degenerative liver
lesions, a frank effect, occurred in cats, rats, mice, rabbits, and
guinea pigs that were exposed to 1.5 mg/m3 Aroclor 1254 vapor for
7 h/day, 5 days/week for 213 days (Treon et al. 1956). This PEL is
plotted on Figs. 2.1 and 2.4. Histologic effects were not produced in
those species exposed to Aroclor 1242 (1.9 mg/m3. 7 h/day, 5 days/week
for 214 days; 8.6 mg/m3, 7 h/day, 5 days/week for 24 days). The higher
-------
18 Section 2
NOAEL of 8.6 mg/m3 for intermediate-duration inhalation exposure is
plotted on Fig. 2.4. Since the FEL for Aroclor 1254 is lower than the
NOAEL for Aroclor 1242, a minimal risk level cannot be derived for
Aroclors as a class.
Developmental toxicity. Pertinent data regarding developmental
effects of PCBs via inhalation exposure in animals were not located in
the available literature. A report of slightly reduced birth weight and
gestational age in infants born to mothers with occupational exposure to
Aroclors (Taylor et al. 1984) is inconclusive and lacks monitoring data.
Reproductive toxicity. Pertinent data regarding reproductive
effects of PCBs via inhalation exposure in humans or animals are not
available.
Genotoxicity. The PCBs have produced generally negative results in
in vivo and in vitro genotoxicity assays (Sect. 4.3.5 on genotoxicity in
toxicological data section).
Carcinogenicity. Occupational studies (Brown 1986, Bertazzi et al.
1987) provide inadequate but suggestive evidence for carcinogenicity of
PCBs by the inhalation route (see Sect. 4.3.6.1). Data regarding the
carcinogenicity of inhaled PCBs in animals are not available.
2.2.1.2 Oral
Lethality and decreased longevity. Data regarding oral exposure
levels that produce death in humans were not available. Single-dose oral
LD5QS for PCBs have been reported for rats and mink. The lowest values
are 750 mg/kg for Aroclor 1221 in mink (Aulerich and Ringer 1977) and
1,010 mg/kg for Aroclor 1254 in rats (Garthoff et al. 1981). These FELs
are plotted on Figs. 2.2 and 2.5 for lethality due to acute oral
exposure.
In mice fed diets containing 1,000 ppm Aroclor 1254 for 14 days, 3
of 5 died by day 15 (Sanders et al. 1974). No mice fed diets containing
250 ppm Aroclor 1254 for 14 days died. Thus, 250 ppm is a NOAEL, and
1,000 ppm is a FEL for lethality in mice for short-term oral exposure.
Assuming that a mouse consumes a daily amount of food equal to 13% of
its body weight (EPA 1986a). the NOAEL is equivalent to 32.5 mgAg/day,
and the FEL is equivalent to 130 mg/kg/day. These levels are plotted on
Figs. 2.2 and 2.5 for lethality for acute oral exposure. Hornshaw et al.
(1986) determined LCSQs of Aroclor 1254 for dietary exposure in mink to
be 79-84 ppm for 28 days and 47-49 ppm for 28 days followed by a 7-day
withdrawal period. In mink fed Aroclor 1254 for 9 months, the LC50 was
6.65 ppm (Ringer et al. 1981). Assuming that mink consume 150 g of feed
per day and weigh 800 g (Bleavins et al. 1980), 47 ppm is equivalent to
an U>50 of 8.8 mg/kg/day (see Fig. 2.2), and 6.65 ppm is equivalent to
an LD50 of 1.25 mg/kg/day. This FEL is plotted on Figs. 2.2 and 2.5 for
intermediate exposure.
Reduced survival occurred in rats fed diets containing 225 ppm
Aroclor 1254 for 104 weeks (NCI 1978). Assuming that rats consume the
equivalent of 5% of their body weight per day in food (EPA 1986a) , then
1.25 mgAg/day represents a FEL for chronic oral exposure in rats (see
Figs. 2.2 and 2.5). NOAELs for Increased mortality were not identified
in these studies.
-------
Health Effects Summary 19
Systemic/target organ toxicity. The liver and cutaneous tissues
are primary targets of PCB toxicity in orally exposed animals.
Rats were fed diets containing 0, 4, 8 or 16 ppm Aroclor 1254 for
4 days (Carter 1985); relative liver weights were significantly
increased at >8 ppm and serum levels of HDL cholesterol were
significantly increased at 16 ppm. The 8-ppm and 16-ppm concentrations,
which correspond to 0.4 and 0.8 mg/kg/day, respectively, if rat food
consumption is assumed to be 5% of body weight per day, represent a
NOAEL and LOAEL for acute oral exposure (see Figs. 2.2 and 2.5). The
NOAEL is the basis for the minimal risk level for acute oral exposure
(Fig. 2.5).
In intermediate-duration studies, hepatic microsomal enzyme
activities were increased in rats treated with diet concentrations of
0.5, 5, or 50 ppm Aroclors 1242, 1248, 1254, or 1260 for 4 weeks
(Litterst et al. 1972). Dietary exposure to 5 ppm Aroclor 1242 for 2 to
6 months produced increased liver lipid content in rats (Bruckner et al
1974) and >20 ppm Aroclor 1254, or 1260 for 28 days (Chu et al. 1977) or
8 months (Kimbrough et al. 1972) produced frank degenerative liver
alterations in rats. Dietary concentrations of 0.5 ppm Aroclors 1242
1248, 1254, and 1260 and 5 ppm Aroclor 1242, therefore, represent the
highest NOAEL and lowest LOAEL, respectively, for intermediate-duration
hepatic effects in rats. Assuming that rats consume 5% of their body
weight in food per day, the NOAEL and LOAEL provided 0.025 and 0 25
ngAg/day, respectively (see Figs. 2.2 and 2.5).
Two monkeys that died from dietary exposure to 2.5 or 5.0 ppm
Aroclor 1248 for 173 or 310 days, respectively, had frank liver lesions
(Barsotti et al. 1976). Although this study is limited by the number of
animals, other studies with monkeys corroborate these FELs, as chloracne
and gastric lesions were also associated with intermediate-duration
exposure to 2.5 or 5.0 ppm Aroclor 1248 (Barsotti and Allen 1975
Barsotti et al. 1976, Thomas and Hinsdill 1978). The lowest monkey FEL
(2.5 ppm) is equivalent to 0.105 mg/kg/day (see Figs. 2.2 and 2.5) if it
is assumed that monkey food consumption is 4.2% of body weight per day
(EPA 1986a).
Chronic feeding studies with rats (NCI 1978, Morgan et al. 1981
Ward 1985, Norback and Veltman 1985, Kimbrough et al. 1975), conducted
at concentrations (>20 ppm) that were higher than the lowest FELs in the
intermediate-duration monkey studies, did not produce degenerative liver
lesions but did produce preneoplastic and proliferative liver lesions.
Chronic (12 to 16 month) feeding studies were conducted with 2.5 and
5.0 ppm Aroclor 1248 in monkeys (Barsotti and Allen 1975, Barsotti et
al. 1976), but skin lesions and other effects (as indicated above and in
subsequent sections) occurred after several months of exposure.
Therefore, it is inappropriate to identify effect levels for systemic
effects resulting from chronic oral exposure because of the types of
liver lesions (preneoplastic) in rats and the short latency for
cutaneous and other effects in monkeys.
Developmental toxicity. Slight effects on birth weight, head
circumference, gestational age and/or neonatal behavior have been
reported in infants of mothers who were consumers of PCB-contaminated
fish (Fein 1984; Fein et al. 1984; Jacobson et al. 1984a, 1984b, 1985)
-------
20 Section 2
and infants of mothers who had no known specific source of FCB exposure
(Rogan et al. 1986, 1987). Although these studies suggest an association
with PCB exposure, these effects cannot conclusively be attributed to
PCBs because of potential and documented exposure to other chemicals,
inconsistency between studies, and other limitations discussed in
Sect. 4.3.3.2.
Collins and Capen (1980a) fed diets containing Aroclor 1254 at 0,
50, or 500 ppm to female rats during gestation and lactation.
Significantly (P < 0.001) reduced litter size occurred at 500 ppm. At
both 50 and 500 ppm, the neonates and weanlings had ultrastructural
lesions in the thyroid follicular cells and reduced serum levels of
thyroid hormone. Thus, 50 ppm is the LOAEL for fetotoxicity due to oral
exposure in rats. Assuming that a rat consumes a daily amount of food
equal to 5% of its body weight (EPA 1986a), 50 ppm is equivalent to
2.5 mg/kg/day. The LOAEL is indicated on Figs. 2.2 and 2.5 for
developmental toxicity in rats.
Gestational exposure to Aroclor 1254 by gavage produced fetotoxic
effects in rabbits exposed on days 1-28 at doses >12.5 mg/kg/day but not
<10 mgAg/day (Villeneuve et al. 1971). The dose of 10 mgAg/day,
therefore, represents a NOAEL for developmental effects in rabbits (see
Figs. 2.2 and 2.5, acute exposure). The dose of 12.5 mg/kg/day
represents a FEL for developmental effects in rabbits because it
produced fetal deaths.
Haake et al. (1987) reported that treatment of pregnant C57BL/6
mice with Aroclor 1254 by gavage at 244 mg/kg on day 9 of gestation did
not result in any fetuses with cleft palate. This dose is plotted on
Figs. 2.2 and 2.5 as a NOAEL for developmental toxicity in mice.
Monkeys that were fed diets containing 1.0 ppm of Aroclor 1016 for
approximately 7 months prior to mating and during pregnancy (total
duration 87 ± 9 weeks) delivered infants with reduced birth weights, but
this effect did not occur at 0.25 ppm (Barsotti and Van Miller 1984).
Assuming that a monkey consumes a daily amount of food equal to 4.2% of
its body weight, the daily dosages in the 1.0 ppm (LOAEL) and 0.25 ppm
(NOAEL) groups were 0.04 and 0.0105 mg/kg/day, respectively. The NOAEL
serves as the basis for the minimal risk level for intermediate and
chronic oral exposure as derived by EPA (1988a). Fetal mortality, a
frank effect, occurred at >2.5-ppm (0.1-mg/kg/day) dietary
concentrations of Aroclor 1248 in other studies with monkeys (Allen and
Barsotti 1976; Allen et al. 1979, 1980).
Reproductive toxicity. There are no studies regarding reproductive
effects of PCBs in humans. Diets that provided >2 ppm of Aroclor 1254
for 4 months prior to mating and during gestation were lethal to fetuses
and caused reproductive failure in mink (Aulerich and Ringer 1977,
Bleavins et al. 1980). Assuming that mink consume 150 g of feed per day
and weigh 800 g (Bleavins et al. 1980), then the 2-ppm FEL provided
0.38 mgAg/day (see Figs. 2.2 and 2.5).
Reduced litter sizes occurred at Aroclor 1254 dietary
concentrations of >20 ppm but not <5 ppm in one- and two-generation
reproduction studies with rats (Linder et al. 1974). The dietary
concentrations of 5 ppm (NOAEL) and 20 ppm (FEL) provided 0.25 and
-------
Health Effects Summary 21
1 mg/kg/day, respectively, if rat food consumption is assumed to be 5%
of body weight per day (EPA 1986a). These levels are plotted on
Figs. 2.2 and 2.5 for reproductive effects of intermediate oral exposure
in rats.
Genotoxicity. The PCBs have produced generally negative results in
in vivo and in vitro genotoxicity tests (Sect. 4.3.5 on genotoxicity in
toxicological data section).
Carcinogenicity. EPA (1988a) used the Norback and Weltman (1985)
study as the basis for a quantitative carcinogenicity risk assessment
for PCBs. The dietary level of 100 ppm Aroclor 1260 was converted to an
intake of 5 mg/kg/day by assuming that a rat consumes food equal to 5%
of its body weight per day. This dosage was converted to a TWA dosage of
3.45 mg/kg/day (see Fig. 2.5) to reflect the fact that rats received
100 ppm for 16 months, 50 ppm for 8 months, and 0 ppm for the last
5 months. The rat dosage was converted to an equivalent human dose of
0.59 mg/kg/day on the basis of relative body surface areas. Incidences
of trabecular carcinomas, adenocarcinomas, and neoplastic nodules in the
liver were combined to produce total incidences of 45/47 in treated
females and 1/49 in controls. Using these data, EPA (1988a) calculated a
human q.* of 7.7 (mg/kg/day)-1. Dosages corresponding to risk levels of
10-*. 16-5 10'6, and 10'7 are 1.3 x 10'5, 1.3 x 10'5, 1.3 x 10'7, and
1.3 x 10'8 mgAg/day, respectively. The 10'4 to 10'6 risk levels are
indicated on Fig. 2.5. Aroclor 1260 is assumed to be representative of
all PCB mixtures because there is no information regarding which
constituents of any PCB mixture might be carcinogenic; therefore, the
potency estimate for Aroclor 1260 applies to all PCB mixtures (EPA
1988b).
2.2.1.3 Dermal
Occupational exposure to PCBs is considered to be by the inhalation
route in this profile, since air levels are commonly monitored in the
workplace. It is clear, however, that under occupational conditions
dermal exposure would also occur. This was recognized by ACGIH (1986)
when a skin notation was placed with the TLV. Dermal adsorption and
exposure can occur from contact of the skin with the vapors of PCB as
well as actual dermal contact with the compound or from contact with
dust or surfaces to which the PCBs are absorbed. Although it is realized
that dermal exposure may be a major route of exposure in the
occupational setting, quantitation of the relative contribution to body
burden of absorbed PCBs from the inhalation and dermal routes is not
possible for most studies. The study of Maroni et al. (1981a,b) permits
some quantitation of dermal exposure, as discussed under systemic/target
organ toxicity below.
Lethality and decreased longevity. Human data are not available.
Median lethal doses for single dermal applications of PCBs to rabbits
ranged from >1,269 mg/kg for Aroclors 1242 and 1248 to <3.169 mg/kg for
Aroclor 1221 (Fishbein 1974). As only ranges of median lethal doses were
reported, the lowest dose (1,269 mg/kg) is indicated on Figs. 2.3
and 2.6.
-------
22 Section 2
Systemic/target organ toxicity. Occupational exposure to PCBs
involves dermal contact, but, for reasons discussed previously,
occupational exposure data were discussed primarily under inhalation
exposure.
The study of capacitor workers by Naroni et al. (1981a,b) indicated
that dermal exposure to PCBs at 2-28 jig/cm2 of skin (on the hands) was
not associated with clear evidence of liver disease, but may have been
associated with liver enzyme induction in some of the workers. Assuming
a total surface area for the hands of 910 cm2 (Hawley 1985) and body
weight of 70 kg, the dermal exposure would have been 0.026-0.364
mg/kg/day. Because the workers were also exposed to PCBs by inhalation
(48-275 pg/m3), and because interpretation of the study is confounded by
the lack of a control group, the dermal exposure range is not plotted on
Figs. 2.3 and 2.6.
Dermal application of Aroclor 1260 to rabbits on 5 days/week at a
dose of 118 mg/day for 38 days (27 total applications) produced
degenerative lesions of the liver and kidneys, increased fecal porphyrin
elimination, and hyperplasia and hyperkeratosis of the follicular and
epidermal epithelium (Vos and Beems 1971). As body weight appeared to be
approximately 2.7 kg (Vos and Beems 1971), the PEL of 118 mg/day is
equal to a dose of 43.7 mg/kg/day (see Figs. 2.3 and 2.6).
Developmental and reproductive tozicity. Pertinent data regarding
developmental and reproductive effects of dermal exposure to PCBs were
not located in the available literature.
Genotoxicity. The PCBs have produced generally negative results it.
in vivo and in vitro genotoxicity tests (Sect. 4.3.5 on genotoxicity in
toxicological data section).
Carcinogenicity. Occupational exposure to PCBs, which involves
inhalation as well as dermal exposure, provides inadequate evidence of
carcinogenicity in humans (Sect. 4.3.6 on carcinogenicity in
toxicological data section). In two-stage carcinogenesis studies with
mouse skin, Aroclor 1254 did not produce evidence of promoter or
complete carcinogen activity and was not tested adequately for initiator
activity (Sect. 4.3.6.3 on carcinogenicity of dermal exposure in
toxicological data section).
2.2.2 Biological Monitoring as a Measure of Exposure and Effects
2.2.2.1 Exposure
PCBs are pervasive environmental contaminants that are found in
body tissues and fluids of the general population. Because they are
lipophilic, PCBs are preferentially stored in adipose tissue and are
present in serum and human milk. Serum and adipose PCB levels are
indicators of exposure, but may not provide accurate estimations of
exposure or body burden because the concentration of PCBs in serum
varies with the concentration of lipids in serum and variations in
procedure and methods of data reporting may preclude interlaboratory
comparison (Kimbrough 1987a).
-------
Heal eh Effects Summary 23
Concentrations of PCBs in human adipose tissue and milk fat are 100
to 200 times higher than in serum (Kimbrough 1987a). Average PCB levels
below 2 ppm in milk fat and 100 ppb in whole milk have normally been
found, and the fat concentration in human milk averages 2 5-4 5% (Jensen
1983, Jensen et al. 1980. Rogan et al. 1987).
In the National Human Adipose Tissue Survey (NHATS), 46 composite
adipose tissue samples collected during surgical procedures or during
autopsies during fiscal year 1982 were analyzed for organochlorine
compounds (EPA 1986b). Of the 46 samples, 83% contained PCBs as follows:
22% contained trichlorobiphenyl, 53% contained tetrachlorobiphenyl, 73%
contained pentachlorobiphenyl, 73% contained hexachlorobiphenyl, 53%
contained heptachlorobiphenyl, 40% contained octachlorobiphenyl, 13%
contained nonachlorobiphenyl, and 7% contained decachlorobiphenyl.
Statistical analyses for baseline estimates and time trends for PCBs in
human adipose tissue in the NHATS for 1970-1983 have been performed
(Lucas 1982, EPA 1985e). These analyses indicate that the estimated
percentage of individuals with PCB levels >3 ppm increased to a peak of
approximately 10% in 1977 and decreased steadily to near zero by 1983.
The percentage of individuals having PCB levels >1 ppm decreased
steadily from a high value near 50% in 1972 to a low value near 9% in
1983. Although these data indicate that PCB amounts are decreasing, the
percentage of individuals with detectable levels (approximately 1 ppm)
increased from approximately 85% in 1972 to nearly 100% in 1983. The
percentage of people who had PCB levels >1 ppm increased with age and
was greater in males than in females, but there was no significant
difference between races. The Northeast Census Region historically
(i.e., in the middle 1970s) had the greatest percentage of people with
PCB levels >1 ppm, but, in recent years, the difference between the
northeast and other regions no longer exists.
Anderson (1985) discussed the use of adipose tissue biopsy in
assessing human exposure to PCBs. Because adipose tissue is the primary
storage site of PCBs, adipose tissue samples have been the preferred
biological specimen. Analysis of PCBs in adipose tissue provides a
direct measure of body burden, but has disadvantages over analysis of
serum levels because collection of samples is invasive and time-
consuming. Based on data that adipose tissue levels of PBBs
(polybrominated biphenyls) and DDT are directly correlated with serum
levels of PBB and DDT, it can be predicted that PCB adipose levels will
also correlate with serum levels. Anderson (1985) recommended that
whenever an adipose tissue sample is obtained at biopsy, a paired serum
sample should be collected and the two tissues be analyzed for PCBs.
Once the correlation is characterized, blood samples may become the
preferred choice for monitoring, unless identification of low exposures
is required.
Wolff (1985) reported data on blood levels of PCBs in workers in
relation to exposure levels (Table 2.1) and blood and adipose tissue
levels of PCBs in workers in relation to duration of employment (Table
2.2). Generally, higher exposure levels result in higher blood and
adipose tissue levels of PCBs, but because PCBs accumulate in the body,
exposure duration is at least as important as exposure level.
-------
24 Section 2
Table 2.1. PCB lereb in blood of exposed workers
(Aroclors 1016, 1242, 1248)
Air levels
(mg/mj)
0.3-2
0.05-0.275
0-0.26
0.1-1
Blood levels
(ng/mL)
Mean High
1,060 3,500
440 1,400
130 407
355 3,330
149 1,500
89 370
118 2,530
48 604
N
19 -Inside""
14 "Outside""
60
26 High exposed
55 Low exposed
140 Never exposed
110 High exposed
180 Other
"Workers who were exposed inside or outside the
impregnation room.
Source: Wolff 1985.
Table 2.2. PCB blood tereb (Aroclor 1254) and duration of exposure
Mean duration
of employment
(years)
16±8
17
3.8
4.3
Mean blood
concentration
(ng/mL)
24"
6*
33'
14'
12'
yv
258
32
86
15
19
Mean adipose
concentration
(Mg/g)
17
4
5.6
1.4
1.3
N
53
8
36
5
9
"Persons with more than 5 years employment; geometric means;
geometric mean of S3 plasma samples which matched the adipose
samples was 54 ng/mL.
^Persons with less than 5 years employment; geometric means.
Tenons exposed.
^Persons nominally exposed.
'Nonexposed.
Source: Wolff 198S.
-------
Health Effaces Summary 25
Kreiss (1985) reviewed available data, including unpublished
Centers for Disease Control (CDC) data, for serum PCB concentrations in
U.S. populations without occupational exposures for 1968-1983. These
data and more recent data of Sahl et al. (1985a,b) and the Massachusetts
Department of Public Health (1987) (i.e., the New Bedford Study) are
summarized in Table 2.3. Mean serum levels were usually between 4 and 8
ng/mL, with 95% of the individuals having concentrations <20 ng/mL
(Kreiss 1985). Cross-sectional data concerning PCB levels in a
representative sample of the U.S. population are not available because
the various groups were monitored during investigations of pesticide
residues, food chain contamination, hazardous waste sites, and
occupational exposure in which a nonexposed control group was necessary.
Mean serum PCB levels in some populations that consumed contaminated
fish are several times higher than mean levels in populations that did
not consume contaminated fish (Table 2.3). The mean PCB levels in these
studies approach those associated with occupational exposure (Table
2.4), but are within the range of the general population groups.
Interpretation of the data in Tables 2.3 and 2.4 is complicated by
differences in analytical methodology and methods of population
selection and data reporting (Kreiss 1985).
PCB levels in adipose tissue and in human milk fat are 100 to 200
times higher than serum levels (Kimbrough 1987a). PCB concentrations
averaged 1.5 ppm in the breast milk of 1,057 women in Michigan (Wickizer
et al. 1981).
2.2.2.2 Effects
Several studies of general population subjects attempted to
correlate serum PCB levels with health indices. Baker et al. (1980)
found that plasma triglyceride levels increased significantly with serum
PCB concentrations in residents of Bloomington, Indiana, including
workers occupationally exposed to PCBs. Chloracne or systemic symptoms
of PCB toxicity were not noted, and there were no significant
correlations between PCB levels and hematologic, hepatic, or renal
function indices. Kreiss et al. (1981) reported that serum PCB levels
were positively associated with serum cholesterol levels, gamma-glutamyl
transpeptidase (GGTP) levels, and measured blood pressure in residents
of Triana, Alabama, that were exposed via consumption of contaminated
fish. Rates of borderline and definite hypertension were 30% higher than
those expected on the basis of national rates. The associations in the
above studies were independent of predictors of PCB levels such as age,
sex, and/or consumption of alcohol and fish. The hypertension and other
effects in the Kreiss et al. (1980) study cannot be attributed solely to
PCBs because the strongest correlation was between log PCB and log DDT
serum levels. Low and moderate serum levels of PCBs did not appear to be
associated with increased blood pressure In residents of New Bedford,
Mass.. who were exposed via consumption of contaminated seafood.
Steinberg et al. (1986) determined chat five serum analytes
(l-glucuronidase, 5'-nucleotidase. triglycerides, cholesterol, and total
bilirubin) correlated positively and significantly with log
concentrations of serum total PCBs in residents who lived or worked in
the vicinity of an electrical manufacturing plant. Aroclor 1260 was
-------
10
TaMtU.
PCB
PCBa and hi
(A
Area and
sampling method
Charleston County, S.C..
Lake Michigan random
aon-Tub-eaten
Bloomington, Ind..
volunteers and controls
Michigan PBB cohort
Billings. Moot, random
Franklin, Idaho,
volunteers
Random uneiposed
workers
Newton. Kans..
1 «k» Michigan random
non-fbh-ealen
Canton. Mass.. volunteers
Old Forge. Pa..
volunteers
Number of
subjects
616
29
110
1.631
17
IOS
19
7
418
10
138
em
Year
1968
1973
1977
1978-79
1979
1979
1979
1979
1980
1980
1981
PCB level. ng/mL
Arithmetic 95%
Arithmetic Geometric standard Confidence
mean mean, median" deviation interval
49
17.3 15*
18.8 - 10.8 17-21
7.7 6.4
7.5 5.8 6.8 4-11
- - - -
12 - - -
4.9 4.2 3.1 2-8
6.6"
7.1 5.2 5.2 3-11
36 - - ~
Range
0-29
-------
TaUt 2J (ttttdMMl)
Are* ud
iaiiiinlinr* «***K««I
Jcffenao. Ohio.
Fairmont. W. Vt.
Norwood. Man.,
volunteen
Ua Angelee-Loot Beech,
Calif., work force*
1 *ke Midiiaan
volunteer apottfithen
Triana. Afau.
votunteen
Lake Michigan
voiutecr tponfiaben
New Bedford. Mam.
nodon ample
New Bedford. Mae*..
kaowo exposure to
Number of
tubjecu
59
40
990
738
90
458
S72
840
110
Arithmetic
Year mean
1983 5.8
1983 6 7
1983 4.9
1982-84 5
1973 717
1979 22.2
1980
1981-82 5.84
1981-82 13.34
PCB level. ng/mL
Arithmetic
Geometric Mandard
mean, median" deviation
4.4 65
5.0 5.3
4.2 3.5
4" 4.37
56-
17.2 22.3
21.4"
3.88" 7 78
9.48" 14.02
95%
Confidence
interval Range
4-8 1-45
5-8 1-23
4-6 2-30
-------
28 Section 2
Table 2.4. Sena PCB cawoKnti
PCB levels. og/mL
Facility
Railway car maintenance
Capacitor plant
Capacitor plant
Capacitor plant
Capacitor plant
Public utility
Private utility
Utility
Number of
subjects
86
34
290
80
221
14
25
I.OS8
Arithmetic
mean
33.4
394"
.24*
48C
342fl
-
-
4
Geometric
mean
67*
21*
119*
25.3'
24*
24'
22*
29*
3"
95%
Confidence
interval
234-554
98-150*
38-58'
:
15-39*
16-35*
17-25*
20-43'
3.65*
Range
10-312
trace- 1.700
6-2.530*
1-546*
41-1.319
1-3.330*
1-250*
5-52*
7-24'
9-48*
7-250*
-------
Health Effects Summary 29
significantly and positively correlated with several of the analytes,
but Aroclor 1242 was correlated significantly and negatively only with
HDL-cholesterol.
Umbilical cord serum levels of PCBs have been correlated with
reduced birth weight and size, shorter gestation, and neonatal
behavioral effects in a few reports (Fein 1984; Fein et al. 1984;
Jacobson et al. 1984a, 1985; Rogan et al. 1986). Although increased
levels of PCBs in cord blood may be predictors of these kinds of
effects, the effects are not well validated and not attributable solely
to PCBs. Cord serum levels associated with these effects are reported in
Sect. 4.3.3 (developmental toxicity in toxicological data section).
Positive correlations between PCBs in blood and levels of
triglycerides and liver-associated enzymes have been reported in workers
with occupational exposure to PCBs (Baker et al. 1980, Ouv et al. 1976,
Fischbein et al. 1979, Haroni et al. 1981b, Chase et al. 1982, Smith et
al. 1982, Fischbein 1985, Lawton et al. 1985, Emmett 1985, Drill et al.
1981, Kreiss 1985). The associations between blood PCBs and
triglycerides should be regarded as equivocal because of partitioning
phenomena, as levels of PCBs in serum appear to be determined by serum
lipid content. Evaluation of associations between serum PCBs and liver-
associated enzymes is complicated by inconsistent and inconclusive data
and lack of correction for confounding variables such as alcohol
consumption. Indicators of possible liver enzyme induction (e.g., GGPT)
are most commonly associated with PCB levels, and associations with
indicators of possible hepatocellular damage (e.g., SCOT, SGPT) have
been demonstrated only in occupationally exposed groups with higher
ranges of PCB levels (Kreiss 1985). The clinical significance of the
alterations in liver-associated enzymes is uncertain, as the increases
may be nonspecific, and indices of obstructive liver disorders have not
been demonstrated even in occupationally exposed groups.
Maroni et al. (1981b) examined the health condition and PCB blood
levels of 80 capacitor manufacturing and testing plant workers who were
exposed to PCBs (42% mean chlorine content) for many years. Sixteen of
the workers had asymptomatic hepatic involvement as determined by
hepatomegaly (12 workers) and serum enzyme elevations (AST, ALT, GGTP,
SCOT and/or SPCH). A significant positive association was found between
the prevalence of hepatic involvement and blood PCB concentrations,
particularly trichlorobiphenyl blood concentrations (X^ trend,
P < 0.001. 0.001 and 0.05 for total chlorobiphenyls, trichlorobiphenyls,
and pentachlorobiphenyls, respectively). Mean blood concentrations of
chlorobiphenyls, trichlorobiphenyls, and pentachlorobiphenyls were
significantly higher in the workers with hepatic involvement compared to
the workers without abnormal findings (Student's t-test, P < 0.001,
0.001 and 0.01, respectively, for the three classes of chlorinated
biphenyls); mean trichlorobiphenyl concentrations were 215 MgAg (range
77-407 pg/kg) in the workers with abnormal liver findings and 92 j*g/kg
(range 13-345 Mg/kg) in those without abnormal liver findings. The
authors suggested that trichlorobiphenyls may reflect current PCB
exposure levels more closely than pentachlorobiphenyls. There were no
significant differences in age or duration of exposure between the
workers with and without abnormal liver findings. Evaluation of the
hepatic findings in this study is complicated by the small number of
-------
30 Section 2
cases, but the enzyme alterations were mild and the prevalence and
severity of the hepatic effects do not appear to be associated with
duration of exposure. Unrelated health problems that may have
contributed to the hepatic effects were described in three of the
workers.
2.2.3 Environmental Levels as Indicators of Exposure and Effects
2.2.3.1 Levels found in the environment
The purpose of this subsection is to summarize available data that
suggest that levels of FCBs found in environmental media (primarily
soil, drinking water, and food) (see Sect. 7.2) are associated with
significant human exposure and/or effects. Schwartz et al. (1983) found
a significant positive correlation (? < 0.001) between fish consumption
measures and PCB levels in maternal serum and milk. The specific PCBs
present were not correlated with the various Aroclor mixtures. From
their data, Schwartz et al. (1983) determined that serum PCB levels
increase by 0.15 ng/mL and milk levels increase by 0.12 ng/g for every
0.45 kg of PCB-contaminated fish consumed, but the rate of fish
consumption by the subjects in the study was not stated. Humphrey (1976)
reported mean blood PCB levels of 0.073 ppm in 105 people whose annual
consumption of Lake Michigan fish equaled or exceeded 24 Ib. The
estimated intake of PCBs by 82% of these people ranged from 0.49 to 3.94
Mg PCB/kg/day and averaged 1.7 /ig/kg/day. Drotman et al. (1983) found a
positive correlation between the PCB concentration in human breast milk
and the number of contaminated eggs consumed by lactating women. The
same form was the source of the eggs in this study, but representative
concentrations of PCBs were not reported. As indicated in Sect. 7.2.4.1,
however, the average concentration of Aroclor residues in contaminated
eggs in 1970-1976 was 0.072 ppm.
2.2.3.2 Human exposure potential
The purpose of this subsection is to discuss the chemical-specific
issues Involved in human exposure of PCBs from water, soil, and food.
Experimental monitoring data have shown that PCB concentrations are
higher in sediment and suspended matter than in the associated water
column, and this is In agreement with the high soil adsorption constants
for PCBs. The partitioning between suspended matter and water will be
isomer specific and should correlate with the octanol/water partition
coefficient of individual isomers. Thus, lover chlorinated PCBs should
have a greater tendency to partition to the water than higher
chlorinated PCBs (see Sect. 6.3.1 on transport and partitioning in
environmental media). This implies that human exposure to the higher
chlorinated isomers from whole water (water + sediment) will be greater
than from settled water. Therefore, all other factors being equal, the
human exposure potential to higher chlorinated PCBs from contaminated
waters may tend to increase as exposure to sediment and suspended matter
increases.
In general, PCBs are strongly adsorbed in most soils; therefore,
leaching will not generally occur. This implies that the exposure will
be greatest at the point of initial adsorption. In many instances, this
may be at or near the soil surface. The principal route of human
-------
Health Effects Summary 31
exposure to PCBs from a spill in soil at a restricted-access outdoor
site is through inhalation of air (EPA 1987a). Soil ingestion and dermal
contact with soil would not be expected to be significant routes of
exposure at a limited-access site. EPA (1987a) calculated that PCB
levels of 25 ppm in soil would present less than a 1 x 10*7 risk to
people on site who work more than 0.1 km from the actual spill area
(assuming that the spill area is
-------
32 Section 2
2.3.2 Health Effect End Points
2.3.2.1 Introduction and graphic summary
The availability of data for health effects in humans and animals
is depicted on bar graphs in Figs. 2.7 and 2.8, respectively.
The bars of full height indicate that there are data to meet at
least one of the following criteria:
1. For noncancer health end points, one or more studies are available
that meet current scientific standards and are sufficient to define
a range of toxicity from no effect levels (NOAELs) to levels that
cause effects (LOAELs or FELs).
2. For human carcinogenicity, a substance is classified as either a
"known human carcinogen" or "probable human carcinogen" by both EPA
and the International Agency for Research on Cancer (IARC)
(qualitative), and the data are sufficient to derive a cancer
potency factor (quantitative).
3. For animal carcinogenicity, a substance causes a statistically
significant number of tumors in at least one species, and the data
are sufficient to derive a cancer potency factor.
4. There are studies which show that the chemical does not cause this
health effect via this exposure route.
Bars of half height indicate that "some" information for the end
point exists, but does not meet any of these criteria.
The absence of a column indicates that no information exists for
that end point and route.
2.3.2.2 Descriptions of highlights of graphs
Data concerning effects of PCBs in humans that are useful for
quantitative risk assessment are not available. The available data
pertain primarily to intermediate- or chronic-duration occupational
exposures in which the exposures are inadequately monitored and do not
correlate with duration and intensity of exposure. Occupational
exposures to PCBs involve significant dermal exposure, but, as discussed
previously, occupational concentrations are expressed in milligrams per
cubic meter of air (mg/m3), which makes it difficult to determine dermal
doses. For this reason, occupational exposure data were discussed under
inhalation exposure. Children born to mothers who consumed PCB-
contaminated fish had some developmental effects, but the effects cannot
be directly attributed to PCBs; therefore, the bar for developmental
effects due to oral exposure indicates that there are some data.
The toxicity and carcinogenicity of the PCBs in animals by the oral
route are reasonably well characterized. Determination of toxicity
effect levels for chronic oral exposure is precluded by occurrence of
proliferative/neoplastic alterations. Effects of acute oral, inhalation,
and dermal exposures to the PCBs in animals have not been extensively
investigated because concern for effects In humans is centered on
intermediate/chronic-duration oral exposures.
-------
HUMAN DATA
V SUFFICIENT
INFORMATION*
SOME
"INFORMATION
NO
INFORMATION
LETHALITY
ACUTE
INTERMEDIATE CHMOMC DEVELOPMENTAL HBPRODOCTIVe CAHCMOOENICITV
/ TOIKITV TOKICITY
SYSTEMIC TOXICITY
'Data a*tot tor occupational wpo«ir«. which la primarty via Inhalatkx., but dannal exposure to likely to occur.
'Sufficient Information extol* to meet at least one of the criteria for cancer or noncancar and pofnta.
Fig. 2.7. Availability of information on health effects of PCBs (human data).
-------
ANIMAL DATA
C/J
(I
o
ft
»-.
§
SUFFICIENT
INFORMATION*
SOME
INFORMATION
NO
INFORMATION
LETHALITY ACUTi INTENMEMATi CHRONIC DEVELOPMENTAL HtPRODOCTIVt CAHCIMOQENICHY
L / TOXICITY TOXKITV
SYSTEMIC TOXICITY
Sufficient information exists to meet at least one of the criteria for cancer or noncancer end points.
Fig. 2.8. Availability of information on health effects of PCBs (animal data).
-------
Health Effects Summary 35
2.3.2.3 Summary of relevant ongoing research
J.L. Jacobson of Wayne State University is conducting a study
sponsored by the National Institute of Environmental Health Sciences to
evaluate the impact of PCBs on physical, cognitive, and neurological
development in early childhood. The children, examined at age 4, were
exposed to moderate levels of PCBs, or maternal serum PCB levels were
high near the time of birth (NTIS 1987) .
W.J. Rogan of the National Institute of Environmental Health
Sciences is conducting a follow-up study of children exposed to PCBs
through breast milk. The children under study are a cohort of 856 North
Carolina children exposed to relatively low levels of PCBs and a cohort
of 108 children from Taiwan exposed to relatively high levels of PCBs
(NTIS 1987).
2.3.3 Other Information Needed for Human Health Assessment
2.3.3.1 Pharmacokinetics and mechanisms of action
Quantitative data concerning the pharmacokinetics of PCBs following
inhalation and dermal exposure are lacking. Such data could greatly
assist efforts to evaluate health effects resulting from inhalation and
dermal exposure to PCBs. Further studies should be conducted concerning
the distribution of PCBs, especially regarding the distribution of PCBs
in the plasma compared to adipose tissue.
Ongoing studies concerning pharmacokinetics and mechanisms of
action were not located.
2.3.3.2 Monitoring of human biological samples
PCBs can be measured in serum, adipose tissue, and milk. These
measurements can indicate elevated exposure but do not provide
information concerning the route of exposure. Although biological
monitoring is useful for documenting exposures, it has limited
applicability at this time.
Biological monitoring methods indicate body burden of PCBs that
have accumulated over a lifetime. Adequate methods are not available to
distinguish exposure routes, short or intermittent exposures, or low-
level exposures due to the bioaccumulation and slow excretion of PCBs.
The Indiana State Department of Health (population survey in Monroe
County, Indiana) is conducting a study that will provide information on
PCB body burden levels in conjunction with selected health outcomes.
Several smaller studies concerning monitoring of biological samples are
being conducted by the CDC.
2.3.3.3 Environmental considerations
Methodology of sufficient sensitivity and specificity to measure
PCBs in the environment exists; however, various laboratories may not
have access to state-of-the-art equipment.
There are no data on the effect of the environmental matrix or
vehicle on the bioavailability of specific PCBs and PCB mixtures.
-------
36 Section 2
Studies with 2,3,7,8-TCDD indicate that the vehicle may play a
significant role in the relative bioavailability of 2,3,7,8-TCDD and
related compounds (e.g., PCBs) (EPA 1985b).
There appears to be a fairly good understanding of the general
environmental fate and transport of PCBs; however, the environmental
fate and transport at specific sites may vary markedly from one site to
another. Therefore, the environmental fate of PCBs at a specific site
may not be understood very well without considerable additional
information. In terms of the general understanding of environmental fate
and transport, more experimental data are required to understand the
potential importance of photolysis in degrading the more highly
chlorinated PCBs, which are more persistent in the environment. In
addition, a better understanding of the environmental cycling of PCBs is
needed to assess future exposure from current environmental sinks such
as PCBs adsorbed to sediments.
No studies were found that involve the environmental interaction of
PCBs with other pollutants.
The U.S. EPA is currently funding studies regarding the
environmental fate and transport of PCBs in the New Bedford Harbor and
the Great Lakes in order to develop data related to this issue.
-------
37
3. CHEMICAL AMD PHYSICAL INFORMATION
3.1 CHEMICAL IDENTITY
Data pertaining to the chemical identity of the Aroclors are listed
in Table 3.1. Aroclors are mixtures of chlorinated biphenyls. The
general chemical structure of chlorinated biphenyls is as follows:
5'
nCI rTCI
(where n and n' may vary from 0 to 5).
The numbering system for the biphenyl structure is also shown
above.
Aroclor products are identified by a four-digit numbering code in
which the first two digits (12) indicate that the parent molecule is
biphenyl and the last two digits indicate the chlorine content by
weight. Thus. Aroclor 1242 is a chlorinated biphenyl mixture with an
average chlorine content of 42%. The exception to this designation
method is Aroclor 1016, which retained the 1016 designation by which it
was known during development (Mieure et al. 1976). Aroclor 1016 is a
mixture that contains primarily mono-, di-, and trichloro isomers and
has an average chlorine percentage (41.5%) that is very similar to
Aroclor 1242.
3.2 PHYSICAL AND CHEMICAL PROPERTIES
Selected physical and chemical properties of the Aroclors are
presented in Table 3.2. Table 3.3 identifies the approximate molecular
composition of the Aroclors.
Data pertaining to the pyrolysis of PCBs, which results in the
formation of polychlorinated dibenzofurans (PCDFs), have been reviewed
(EPA 1988a). Several studies involving pyrolysis of specific PCB isomers
have found that the pyrolysis products include PCDFs, chlorinated
benzenes, naphthalenes, phenyl ethynes, biphenylenes, and hydroxy PCBs.
There appear to be four major paths for production of PCDFs from PCBs:
(1) loss of two ortho chlorines, (2) loss of ortho hydrogen as well as
chlorine, (3) loss of an ortho hydrogen as well as chlorine but
involving a shift of chlorine from the 2 to the 3 position, and (4) loss
of two ortho hydrogens (EPA 1988a). The formation of PCDFs from the
-------
38 Section 3
pyrolysls of PCBs occurred when an electrical transformer in an office
building in Binghamton, New York, accidentally caught fire on
February 5, 1981 (Schecter and Tieman 1985, Tiernan et al. 1985)
-------
Table 3.1. ChMkal Undly oflkt Arodon
Chemical name"
Synonyms
Trade names
Chemical formula
Wuwcsscr line notation'
Chemical structure
Identification Nos.
CAS Registry No.
NIOSH RTECS No.
EPA Hazardous
MI . ^i_ tf
Waste No.
OHM-TADS No
DOT/UN/NA/IMCO
Shipping No.
STCCNo.
Hazardous Substances
Data Bank No.
National Cancer
Institute No
"These are the current
Aroclor 1016
PCB-IOI6
Polychlonnaied
biphenyl with
41 5% Cl
Aroclor*
See Table 3.3
NA
Sec text
12674-1 1-2
TQI 35 1000
3502
8500400
UN23I5
4961666
Unknown
Unknown
Aroclor 1221
PCB-I22I
Polychlonnaied
biphenyl with
21% Cl
Aroclor
Sec Table 3 3
NA
Sec text
II 104-28-2
TQI 352000
3502
8500401
UN23I5
4961666
Unknown
Unknown
Aroclor 1232
PCB-1232
Polychlonnaied
biphenyl with
32% Cl
Aroclor
Sec Table 3 3
NA
Sec text
III4I-I6-S
TQI 354000
3502
8500402
UN23I5
4961666
Unknown
Unknown
Aroclor 1242
PCB-1242
Polychlonnated
biphenyl with
42% Cl
Aroclor
See Table 3 3
NA
Seeiexl
53469-21-9
TQI 356000
3502
8500403
UN23I5
4961666
Unknown
Unknown
Aroclor 1248
PCB-1248
Polychlonnated
biphenyl with
48% Cl
Aroclor
Sec Table 3 3
NA
See text
12672-29-6
TQI 358000
3502
8500404
UN23I5
4961666
Unknown
Unknown
Aroclor I2S4
PC B- 1254
Polychlonnaied
biphenyl with
54% Cl
Aroclor
See Table 3.3
NA
Seeiexl
1 1097-69-1
TQI 360000
3502
8500405
UN23I5
4961666
Unknown
C02664
Aroclor 1260 References
PCB-1260 SANSS 1987
Polychlorinaled
biphenyl with
60% Cl
Chlorodiphenyl
(60% Cl)
Aroclor
Sec Table 3.3
NA
Sec text
1 1096-82-5 SANSS 1987
TQI 362000 SANSS 1987
3502 EPA I980a
8500406 EPA-NIH 1987
UN23IS Chcmlme 1987
4961666 Stone 1981
1822 HSDB 1987
Unknown NCI 1978
chemical names as indexed by ihe Chemical Abstracts Service (CAS)
* Aroclor is the trade name for chlorinated
biphenyls made by
Monsanto
'Wiswcsser line notations are not applicable for mixtures
'Designation prior to May 19. 1980
rj
a-
o
to
*"
to
Q.
T)
CO
»-.
to
K.
H
a
o*
to
a
§
l/J
vO
-------
aase j.*. raystcai asw CMWJSCHI proptnsca 01 rvoa ^
A
Aroclor designation
Molecular weight*
Color
Physical stale
Odor
Melting point, *C
Boiling pout. "C
(distillation range)
Autoignition temperature
Solubility
Water, mg/L
Organic solvents
Density, g/cm' at 25*C
Partition ooefficienl
Log ocunol-watcr
Vapor pressure.
mm Hg at 25*C
Henry's law constant.
atm-m'/mol at 25«C*
Refractive iodei
1016
257.9
Clear
Oil
Unknown
Unknown
325-356
Unknown
0.42
Very soluble
1.33
5.6
4 X IO"4
2.9 X IO'4
1.621 5-1.6235
(25«C)
1221
200.7
Clear
Oil
Unknown
Unknown
275-320
Unknown
0.59 (24°C)
Very soluble
1.15
4.7
6.7 X 10-'
3.5 X IO'1
I.6I7-I6I8
(20°C)
12)2
232.2
Clear
Oil
Unknown
Unknown
290-325
Unknown
Unknown
Very soluble
124
5.1
406 X KT1
Unknown
Unknown
1242
266.5
Clear
Oil
Unknown
Unknown
325-366
Unknown
024
0.34
O.IO(24*C)
Very soluble
1.35
5.6
406X ID'4
5.2 X IO'4
1 627-1.629
(20»C)
1248
299.5
Clear
Oil
Unknown
Unknown
140-375
Unknown
0.054
0.06 (24°C)
Very soluble
141
6.2
4.94 X I0~4
2.8 X 10"'
Unknown
1254
3284
Light yellow
Viscous liquid
Unknown
Unknown
365-390
Unknown
0.012
0.057 (24°C)
Very soluble
1.50
6.5
7.71 X IO'1
2.0 X IO'1
1.6375-1.6415
(25"C)
1260
375.7
Light yellow
Sticky resin
Unknown
Unknown
385-420
Unknown
0.0027
Very soluble
1.58
6.8
405X 10-'
46 X IQ-'
Unknown
0
PI
9
References a
Hulanger el al 1974
Monsanto 1974
Monsanto 1974
Monsanto 1974
Monsanto 1974.
Paris el al. 1978.
Holbfield 1979
EPA 19858
Monsanto 1974
6
Monsanto 1974.
Callahan el al. 1979
c
IARC 1978
-------
Table 3.2 (coMloued)
Aroclor designation
Flash point. "C
(Cleveland open cup)
FUmmability limits
Conversion factors
Air (25°C)*
1016
Unknown
Unknown
1 mg/m1 -
0095 ppm
1221
176
Unknown
1 mg/m1
012 ppm
1232
238
Unknown
1 mg/m'
0 105 ppm
1242
None
Unknown
1 mg/m1 -
0092 ppm
1248
None
Unknown
1 mg/m1 -
008 ppm
1254
None
Unknown
1 mg/m1 -
0.075 ppm
1260
None
Unknown
1 mg/m1 -
0.065 ppm
References
Hubbard 1964
"Average matt from Table 3 3
Tbete log £.. values represent an average value for the major components of the individual Aroclor. Experimental values for the individual components were
obtained from Hansch and Leo 1985
f These Henry's law constants were estimated by dividing the vapor pressure by the water solubility The first water solubility given in this table was used for the
calculation The resulting estimated Henry's law constant is only an average for the entire mixture, the individual chlorobiphenyl isomers may vary significantly from
the average Burkhard el al (1985) estimated the following Henry's law constants (atm-m'/mol) for various Aroclors at 25°C 1221 (2 28 X I0~4). 1242 (3 43 X
IO-*). 1248(4.4 X lO'4), 1254(2.83 X IO'4). 1260(4.15 X 10 «).
rfThe*e air conversion factors were calculated by using the average molecular mass as presented under molecular weight
§
K-
O
to
I
to
»-.
o
to
-------
42 Section 3
laow jj. Approximate moMcuiw coaposmoo 01 rvns
(percent)
Empirical formula
C|2H|0
CI2H9C1
C,2H,C12
C|2fi7Cl}
C|2HgCl4
C12H,C1,
C,2H4CI«
CI2H,C1,
C12H2C1,
C12H,C1,
Average molecular
mass
Aroclor designation
1016
<0.1
1
20
57
21
1
<0.1
ND
ND
ND
257.9
1221
11
51
32
4
2
<0.5
ND
ND
ND
ND
200.7
1232
<0.1
31
24
28
12
4
<0.1
ND
ND
ND
232.2
1242
<0.1
1
16
49
25
8
1
<0.1
ND
ND
266.5
1248
ND"
ND
2
18
40
36
4
ND
ND
ND
299.5
1254
<0. 1
<0.l
0.5
1
21
48
23
6
ND
ND
328.4
1260
ND
ND
ND
ND
1
12
38
41
8
ND
375.7
ND none detected.
Source: Hutanger et al. 1974.
-------
4. TOZICOLOGICAL DATA
4.1 OVERVIEW
Evaluation of che toxicokinetics and toxicity of PCBs is
complicated by the fact that PCBs are mixtures of a variety of different
congeners and impurities, each with its own characteristics. Aroclor
PCBs are the subject of this profile, but toxicokinetic studies often
examined specific congeners, and many toxicological studies used
mixtures of PCBs other than Aroclors, particularly Kanechlors and
Clophens. Kanechlors and Clophens are similar to Aroclors but are
produced in Japan and Germany, respectively, rather than in the United
States, and differ in methods of production, chlorine composition, and
polychlorinated dibenzofuran (PCDF) contamination. The reported range of
PCDFs is 0-2 ppm in Aroclors and 5-20 ppm for Japanese and European PCBs
(Drill et al. 1981). Reference to Kanechlors and Clophens is made
occasionally to support statements made about Aroclors because effects
produced by Aroclors, Kanechlors, and Clophens are generally assumed to
be similar, particularly for mixtures of equivalent chlorine percentages
(Kimbrough 1987a). Non-Aroclor toxicity data are not considered in
detail because of the aforementioned differences in composition and
because reported lowest effect levels are lower for Aroclors than
Kanechlors or Clophens.
The general population is exposed to PCBs primarily by the oral
route (through food, particularly fish). It is possible that indoor air
may be a significant source of PCB exposure. Inhalation and dermal
exposure are the primary routes of occupational exposures, but the
relative contribution of these routes is unknown.
Studies of the absorption of PCBs following oral exposure indicate
that gastrointestinal absorption of most isomers is >90%. The limited
data concerning the absorption of PCBs following inhalation and dermal
exposure indicate that PCBs can be absorbed via these exposure routes,
but the data are not sufficient for quantitative estimates.
Distribution of ingested or injected Aroclors follows a biphasic
pattern. During the first day following dosing, the PCBs distribute to
the liver and muscle tissue. Because of their lipophilicity, the PCBs
are then redistributed to the fat, skin, and other fat-containing
organs. Heavily chlorinated congeners redistribute to adipose tissue to
a greater extent than the less chlorinated congeners, although the type
of chlorine substitution is also a factor.
A number of studies indicate that PCBs can cross the placenta and
locate in the fetus. PCBs also concentrate in milk. Higher PCB levels
may reach the offspring through nursing than through placental transfer.
-------
44 Section 6
The metabolism of PCBs is dependent on the number and position of
chlorine atoms, with lesser chlorinated isomers metabolized more readil,
than more chlorinated isomers. PCB metabolites tend to be 3- or 4-
hydroxy compounds. Evidence suggests that metabolism proceeds through an
arene oxide intermediate except for the 3-hydroxy metabolites, which are
formed by a different pathway involving, at least in part, direct
hydroxylation. The position and degree of chlorination substantially
influence the rate and extent of metabolism. Metabolism is facilitated
by the presence of at least two adjacent unsubstituted ring carbons,
particularly in the 3,4,5 or 3',4',5' positions.
PCBs that are metabolized with more difficulty tend to be excreted
almost exclusively through the biliary route, while the metabolites of
mono-, di-, and trichlorinated isomers are also eliminated through the
urine. Urinary metabolites are in the form of conjugates, including
glucuronides and sulfates. Glutathione conjugates have also been
identified.
Higher chlorinated PCBs tend to persist in the body longer than
lower chlorinated PCBs. For example, biological half-lives in the rat
range from approximately 1 day for 2,2'-dichlorobiphenyl to 460 days for
2,2',4,4',5,5'-hexachlorobiphenyl.
Aroclors appear to have a low order of acute lethality. Data for
non-Aroclor PCB mixtures and specific PCB isomers suggest that mice and
guinea pigs are more sensitive than rats. Aroclors are lethal at much
lower total doses when administered subchronically or chronically than
acutely, indicating that PCBs bioaccumulate to concentrations that are
toxic.
Animal studies have shown that the liver and cutaneous tissues are
the major target organs for Aroclors. Aroclors have also been shown to
produce stomach and thyroid alterations, immunosuppressive effects, and
porphyria in animals. Animals are sensitive to repeated exposures to
Aroclors as a result of rapid bioaccumulation to toxic levels. Monkeys
are particularly sensitive to the toxic effects of Aroclors. Toxic
effects have not been documented in humans who were exposed to Aroclors
via the environment. Occupational exposure to Aroclors has been
associated with reversible skin lesions and subclinical alterations in
serum enzymes that are suggestive of liver enzyme induction and possible
hepatocellular damage.
More serious health effects were observed in humans who consumed
rice oil that had been contaminated with Kaneclors in Japan ("Yusho"
incident) and Taiwan ("Yu Cheng" incident). Although there is an
historical linkage between Yusho and PCBs and some regulatory documents
ascribe health effects from these incidents to PCBs, effects from the
incidents are not reviewed in this report because exposure was to
Kaneclors and because the effects cannot be attributed specifically to
the Kaneclors. The Kaneclors were heated in thermal heat exchangers
before the rice oil contamination and during cooking and contained
relatively high concentrations of PCDFs and polychlorinated quaterphenyl
contaminants. There appears to be general agreement that the PCDF
contaminants, particularly the more potent isomers, contributed
significantly to the health effects observed in the Yusho and Yu Cheng
-------
Toxicologies! Data 45
patients. Please refer to Kuratsune and Shapiro (1984) and Kimbrough
(1987a) for a more complete discussion of this topic.
Aroclors appear to be fetotoxic but not teratogenic in various
species of animals, including rats, mice, rabbits, and monkeys, but the
possibility that contaminants (e.g., PCDFs) may be responsible for the
effects should be recognized. Slight decreases in birth weight,
gestational age, and/or neonatal behavioral performance have been
reported in infants born to mothers who had environmental or
occupational exposure to PCBs. These effects are inconclusive and not
definitely attributable to PCBs.
Oral exposure to Aroclors produced deleterious effects on
reproduction in monkeys, mink, and, at higher doses, rodents.
PCBs have produced generally negative results in in vitro and in
vivo mutagenicity assays.
Feeding studies in laboratory animals demonstrated the
carcinogenicity of several PCB mixtures, but it is not clear which
components of the mixture or metabolites are actually carcinogenic. The
liver is the primary target of PCB carcinogenicity.
4.2 TOXICOKZNETICS
4.2.1 Absorption
4.2.1.1 Inhalation
Human. Inhalation exposure and dermal exposure are the primary
routes of occupational exposure to PCBs, but the relative contribution
of each route has not been discerned (Wolff 1985).
Animal. Six rats were exposed to an aerosol of a PCB mixture
(Pydraul A200, 42% chlorine) at a concentration of 30 g/m^ (0.5 to 3 ^ra
particles) (Benthe et al. 1972). PCB concentrations in the liver after
exposure for 15 min were >50% of the maximum concentration attained
after exposure for 2 h (70 jig/g tissue). These data indicate that the
PCBs were readily absorbed, but the data were not sufficient for more
quantitative estimates of amount or rate of absorption.
4.2.1.2 Oral
Human. The general population is exposed to PCBs primarily by the
oral route (primarily by consumption of contaminated fish). Schwartz et
al. (1983) found elevated levels of PCBs in the serum and breast milk of
women who ate PCB-contaminated fish from Lake Michigan. Humphrey (1976)
reported blood levels of PCBs in people who consumed contaminated sport
fish from Lake Michigan in 1973. Annual consumption of £24 Ib resulted
in a mean blood level of 0.073 ppm (n - 105, s.d. not reported), while
annual consumption of 56 Ib resulted in a mean blood level of 0.020 ppm
(n - 37, s.d. not reported). Blood levels of PCBs in persons who ate no
fish averaged 0.017 ppm (n - 16, s.d. not reported). The estimated
intake of PCBs by 82% of the people who consumed 224 Ib ranged from 0.49
to 3.94 ng PCBAg/day and averaged 1.7 pgAg/day- These studies indicate
that PCBs are absorbed by the gastrointestinal tract, but do not provide
information regarding the extent of absorption.
-------
46 Section 4
Animal. Drill et ml. (1981) and EPA (1985a) reviewed a number of
animal studies indicating that PCBs, including Arodors, are absorbed
readily from the gastrointestinal tract following oral administration.
Albro and Fishbein (1972) examined the absorption of 19 PCB congeners
and unchlorinated biphenyl in rats treated by gavage at doses of 5, SO,
or 100 fflg/kg. Determination of PCBs in faces collected for 4 days
indicated that absorption of all congeners was >90%. Using rhesus
monkeys, Allen et al. (1974a,b) determined over 2-week periods that >90%
of a single oral dose of 1.5 or 3.0 gAg Aroclor 1248 was absorbed.
Bleavins et al. (1984) determined over a period of 5 weeks that European
ferrets absorbed 85.4% of a single dose of [14C]-labeled Aroclor 1254
(0.05 mg) given in food.
In contrast to the above studies, Norback et al. (1978) claimed
that 59.3 to 87% of a single oral dose of 2,4,5,2',4',6*-
hexachlorobiphenyl passed unabsorbed through the intestine of monkeys
during the first week after dosing. It was unclear why relatively little
of this isomer was absorbed. There are no data on the effect of the
environmental matrix or vehicle on the bioavailability of specific PCBs
and PCB mixtures. Studies with 2,3,7,8-TCDD indicate that the vehicle
may play a significant role in the relative bioavailability of 2,3,7,8-
TCDD and related compounds (e.g., PCBs) (EPA 1985b).
4.2.1.3 Dermal
Human. In a study of occupational exposure of electrical workers
to PCBs (Pyralen 3010 and Aplrolio, 42% chlorine content), Maroni et al
(1981b) concluded that absorption of PCBs occurred mainly through the
skin. Quantitive data were not available.
Animal. Single doses of 14C-labeled PCBs (42% and 54% chlorine
content) were applied to the skin of rhesus monkeys (4.1 and 19.3 pg/cm2
42% chlorine) and guinea pigs (4.6 pg/cm2 42% chlorine and 5.2 pg/cm2
54% chlorine) that were lightly clipped of hair (Wester et al. 1983).
The application sites were washed with water and acetone after 24 h, and
radioactivity in the urine was determined during the 28 days (monkeys)
and 16 days (guinea pigs) following dosing. Absorption ranged from
approximately 15-34% of the applied radioactivity in the monkeys and
averaged approximately 33% (42% chlorine) and 56% (54% chlorine) of the
applied radioactivity in the guinea pigs. When 14C-labeled PCB (42%
chlorine) was applied to guinea pig skin and immediately washed with
water and acetone, approximately 59% of the dose was recovered, when
both mixtures were applied to guinea pig skin, left for 24 h and then
washed, approximately 1% of the 42% chlorine concent PCB and 20% of the
54% chlorine content PCB doses were recovered.
Using tritium-labeled PCBs (40% chlorine), Nishizual (1976) found
evidence for dermal absorption of PCBs in rat* via follicular diffusion.
Quantitative data were not provided.
4.2.2 Distribution
4.2.2.1 Inhalation
Human. Wolff et al. (1982b) examined th* relative concentrations
of PCB congeners in plasma and adipose tissue of 26 persons
-------
Toxlco log Leal Data 47
occupatlonally exposed to PCBs (20 to 54% chlorine). Exposure was not
discussed, but it probably included both inhalation and dermal exposure.
The results indicated that PCB congeners with chlorines in both
4 positions were the major components in plasma and adipose tissue. PCBs
with unsubstituted 3,4 positions on at least one ring were observed at
lower concentrations in plasma and adipose. The adipose-plasma partition
ratio calculated for Aroclor 1248 residues was 185, while the partition
ratio for Aroclor 1254 residues was 190. In a more recent study of 173
workers from the same population, adipose-plasma partition ratios of
210, 190. and 200 were determined for Aroclors 1242, 1254, and 1260,
respectively (Brown and Lawton 1984).
Animal. Maximum PCB concentrations in the liver and brain of rats
occurred 2 and 24 h, respectively, after a single 30-min exposure to
30 g/m3 of Pydraul A200 aerosol (42% chlorine) (Benthe et al. 1972).
Concentrations in these tissues subsequently declined, while adipose
concentrations reached a maximum after 48 h.
4.2.2.2 Oral
Human. A number of studies reviewed by EPA (1988a) indicate that
PCBs concentrate in human breast milk. Exposures in these studies were
most likely oral, but may have included both inhalation and dermal
exposure. Wolff (1983) reported the half-life for the decline in
concentration of PCBs (percentage chlorine in compounds not stated) in
breast milk to be 5 to 8 months and found that the concentration of PCBs
in breast milk was 4 to 10 times that in maternal blood. Similar results
were reported by Jacobson et al. (1984b).
Ando et al. (1985) examined the PCB concentration in maternal blood
and milk and the placenta of six Japanese women. They found that the
congeners present were more typical of Kanechlor 500 than Kanechlors
300. 400, or 600. The results indicated that as the chlorine content of
the PCB congeners increased, the correlation between the placental
content of congeners and maternal blood and milk also increased.
PCBs were detected in the umbilical tissues, umbilical blood,
amniotic fluid, and baby's blood from a woman who was occupationally
exposed to Kanechlors 300 and 500 in a capacitor factory (Yakushiji et
al. 1978). PCB levels in these tissues and fluids were considerably less
than in the mother's blood. Maternal serum concentrations of PCBs were
also higher than cord serum concentrations in women who resided in
western Michigan (Jacobson et al. 1984b) and upstate New York (Bush et
al. 1984) (i.e., in regions with easy access to fish from the Great
Lakes).
Kraul and Karlog (1976) determined PCB levels in abdominal fat,
brain, and liver from necropsies completed in 1972 and 1973 in
Copenhagen, Denmark. The ratios of PCB levels were reported as 1:3.5:81
for brain:liver:fat, indicating that adipose was the site of greatest
bioaccumulation of the tissues studied.
Animal. Following absorption. PCBs, including Aroclors, are
distributed in a biphasic manner. The compounds rapidly (minutes to
hours) clear from the blood and accumulate in the liver and muscles
(Drill et al. 1981). PCBs may be translocated from the liver to adipose
tissue for storage or be metabolized in the liver, with metabolites
-------
48 Section 4
excreted in the urine or bile. The accumulation of PCBs in lipophilic
tissues is dependent on the structure-dependent metabolic rates of the
individual congeners.
Muehleback and Bickel (1981) treated rats by gavage with a single
dose of 0.6 or 3.6 mgAg [14C]-2,4,5,2',4',5'-hexachlorobiphenyl. The
rats were examined 1 h, 24 h, 6 weeks, 20 weeks, or 40 weeks after
dosing. The results showed the highest levels of PCBs in muscle, liver,
fat, and skin early in the study. By the end of the study, the highest
PCB levels were found in adipose tissue followed by skin, muscle, and
liver. During the 40-week study period, only 16% of the total dose was
excreted.
Gage and Holm (1976) determined concentrations in abdominal fat of
mice 7 and 21 days after the mice were dosed by gavage with a single
dose (13-165 /jg/mouse) of 1 of 14 PCB congeners. Relatively low levels
(<10 ng/g//ig dose) were found at 7 days for 4,4'-, 3,2',4',6'-, and
2,3,4,2',4*,6'-isomers, with relatively high levels (ilOO ng/g/ug dose)
for 2,4,5,2',4',5-, 4,2'4'6'-, and 2,4,2'4'-PCBs.
Kurachi and Mio (1983a) exposed mice to Kanechlor 400 at 100 mgAg
in the diet for 5 to 20 days. Analysis of tissues at the end of the
feeding period indicated high levels of PCBs in the gonads. High levels
of PCBs were also found in skin, adipose tissue, adrenals, and kidneys.
A second group of mice were kept on the PCB diet for 20 days in a
rotation cage to cause fatigue. Mobilization of fat deposits was
observed with liver PCB levels in fatigued mice 10 times greater than in
mice fed the same diets but allowed to rest.
A number of animal studies have demonstrated that PCB mixtures and
specific congeners and isomers can cross the placental barrier and enter
the fetuses (EPA 1988a). High levels of PCBs accumulate in the mammary
gland where they are secreted in the fat portion of the milk. Masuda et
al. (1979) fed PCBs to pregnant mice through the first 18 days of
gestation and found the highest levels of serum PCBs in offspring 1 to
2 weeks old as compared with 18-day fetuses or with older offspring. In
studies in which monkeys were exposed prior to and during gestation,
signs of PCB-induced intoxication in nursing but not newborn offspring
were observed (Allen and Barsotti 1976, latopoulos et al. 1978). Results
such as these have led some investigators to conclude that transfer
through nursing may account for higher exposure of young than does
placental transfer.
Groups of 24 rhesus monkeys were maintained on diets that provided
Aroclor 1016 at doses of 0, 4.5, or 18.1 mg/kg/day throughout gestation
and a 4-month gestation period (Barsotti and Van Miller 1984). At birth,
the concentrations of the PCBs in the skin of infants were similar to
concentrations in the subcutaneous fat of the mothers. At weaning, the
PCB content In the mesenteric fat of the infants was 4 to 7 times
greater than in the subcutaneous fat of the aethers. Gas chromatographic
patterns showed that the adult adipose tissue did not Include the total
spectrum of peaks observed in the Aroclor 1016 standard, that all of the
peaks observed in the standard occurred in the neonate skin, and that
the peaks In the mesenteric fat at weaning and 4 months after weaning
were qualitatively similar to those in the adult adipose tissue. These
data suggested an inability of the fetus to metabolize and excrete
-------
Toxicological Data 49
certain congeners that are more readily metabolized and eliminated by
adults and older infants.
Bleavins et al. (1984) fed female European ferrets a single dose of
(L*C]'labeled Aroclor 1254 in the diet (0.05 mg) early (day 14) or late
(day 35) in gestation and determined the placental transfer of PCBs.
They found that placental transfer to the kits was 0.01% (per kit) of
the maternal dose when dams were exposed early in gestation and 0.04%
(per kit) when dams were exposed late in gestation. Placental transfer
of PCBs was considerably less than mammary transfer, with the ratio of
placental to mammary transfer at 1 week of lactation 1:15 and 1:7 for
offspring of dams dosed early and late in gestation, respectively.
4.2.2.3 Dermal
Data concerning the distribution of Aroclors following dermal
exposure of humans or animals were not located. Because PCBs are
lipophilic, the compounds should concentrate in adipose tissue
regardless of the route of exposure.
4.2.3 Metabolism
4.2.3.1 Human
2,2',4,4',5,5'-Hexa-CB was the PCB congener found in the highest
concentration in human adipose tissue, while 2,2',4,4',6,6'-hexa-CB was
not detected (Jensen and Sundstrom 1974). As both of these compounds are
found in commercial PCB mixtures and in the environment, the presence of
the 2,2',4,4',5,5'-hexa-CB congener in adipose tissue appears to be
related to resistance to metabolism (EPA 1988a). That this congener is
not metabolized or is minimally metabolized is also indicated by the
finding that the blood concentration of this congener in 17 PCB-poisoned
patients decreased only 10% over 300 to 500 days (Chen et al. 1982). The
measurements began 7 months to a year after the outbreak of poisoning.
The results of in vitro metabolism studies with human liver microsomes
also demonstrate minimal metabolish of this congener (Schnellman et al
1983, 1984).
There were lower concentrations of PCBs with unsubstituted
3,4 positions on at least one of the phenyl rings than PCBs with
substitutions in the 2,4 or 3,4 positions on both rings in the blood and
adipose tissue from capacitor-manufacturing facility workers (Wolff et
al. 1982a).
4.2.3.2 Animal
The metabolism of PCBs has been investigated in numerous studies
with animals and reviewed by EPA (1988a) and Drill et al. (1981). A
variety of substrates have been tested, and the PCBs were usually
administered by the oral or parenteral routes. General findings of these
studies reported by EPA (1988a) are presented below.
Phenolic products are the major PCB metabolites although sulfur-
containing metabolites (e.g., methylsulfones), trans-dihydrodiols,
polyhydroxylated PCBs, and methyl ether derivatives have also been
identified. Although the effects of chlorine substitution patterns on
sites of oxidation have not been studied systematically, EPA (1988a)
suggests the following:
-------
50 Section 4
1. Hydroxylation is favored at Che para position in the least
chlorinated phenyl ring unless this site is sterically hindered
(i.e., 3,5-dichloro substitution).
2. In the lover chlorinated biphenyls, the para position of both
biphenyl rings and carbon atoms that are para to the chloro
substituent are all readily hydroxylated (Sparling et al. 1980).
3. The availability of two vicinal unsubstituted carbon atoms
(particularly CS and C4 in the biphenyl nucleus) also facilitates
oxidative metabolism of the PCB substrate but is not a necessary
requirement for metabolism.
4. As the rate of chlorination increases on both phenyl rings, the
rate of metabolism decreases.
5. The metabolism of specific PCB isomers by different species can
result in considerable variations in metabolite distribution.
PCB metabolites tend to be 3- or 4-hydroxy compounds. The
occurrence of trans-dihydrodiol metabolites suggests that metabolism of
PCBs proceeds through formation of arene oxide intermediates (EPA
1988a). 3-Hydroxybiphenyl appears to be formed by a different mechanism,
at least in part via direct hydroxylation (Billings and McMahon 1978).
Arene oxides are potential electrophiles that have been implicated in
cellular necrosis, mutagenicity, and carcinogenicity. The toxicological
significance of PCB metabolism is unknown, but most studies suggest that
the parent hydrocarbon initiates most of the common toxic responses by
initial binding to the cytosolic receptor protein (EPA 1988a). The role
of metabolism in the genotoxicity of PCBs has not been delineated.
PCB metabolites are usually more polar than the parent compounds
and conjugated with glucuronides or sulfates prior to elimination. Rats
and mice that were exposed to di-, tetra-, or penta-CBs by
intraperitoneal injection or diet eliminated metabolites of glutathione
conjugates and other sulfur-containing compounds (Kurachi 1983, Kurachi
and Mio 1983b).
4.2.4 Excretion
4.2.4.1 Inhalation
Data concerning the excretion of PCBs in humans and animals
following inhalation exposure were not available.
4.2.4.2 Oral
The excretion of PCBs is to a large extent dependent on the
metabolism of PCBs to more polar compounds (EPA 1988a). At equilibrium,
the elimination of PCBs from all tissues will be dependent on the
structure-dependent metabolism rates of the individual PCB congeners.
For example, biological half-lives in the rat range from 1.15 days for
2,2'-dichlorobiphenyl to approximately 460 days for 2,2',4.4',5,5'-
hexachloroblphenyl (Tanabe et al. 1981, Wyss et al. 1986). Metabolites
of the more highly chlorinated congeners are eliminated primarily via
the feces (Goto et al. 1974).
-------
Toxicologleal Data 51
Hunan. Chen et al. (1982) report on the determination of PCBs in
the blood of humans in Taiwan after they consumed rice-bran oil
contaminated with Kanechlor 500 and PCDFs. Blood samples from 17
patients were examined, with 2 to 3 samples taken from each patient 2 to
17 months apart. The results indicated that the tetra- and some penta-
Isomers tend to be eliminated more rapidly than other penta-, hexa-, and
hepta-isomers. Half-lives for the 2,4,5,2',4'- and 2,3,4,3',4'-penta-
isomers in blood were determined to be 9.8 and 8.7 months, respectively.
The data also indicated that two adjacent unsubstituted carbon atoms at
the meta-para positions facilitated metabolism and subsequent
elimination from the blood.
Animal. Hashimoto et al. (1976) examined the excretion of [14C]
PCB compounds given once a week to rats by gavage at a total dose of
6.35 to 7.85 mg/kg over a period of 5 to 50 weeks. The PCBs studied were
predominantly tetra- and hexa-chlorinated isomers. The results indicated
that 1.9 to 4.9% of the dose of tetra-PCBs was excreted in the urine
collected for 7 days after the last dose, with the higher amounts
excreted in rats treated for longer periods. In rats treated with hexa-
PCBs, only 0.6% of the dose was excreted in the urine collected for
7 days after the last dose (treatment was for 5 weeks only). About 47 to
68% of the dose of both tetra- and hexa-isomers was excreted in the
feces, most of which was excreted in 2 days after the last dose.
Mizutani et al. (1977) studied the elimination of tetra-CB isomers
in mice fed diets containing a single isomer at 10 ppm for 20 days.
Biological half-lives for the individual isomers were 0.9, 9.2, 3.4,
0.9, and 2.1 days for 2,3,2',3'-; 2.4,2',4'-; 2,5,2',5'-; 3,4,3'4'-; and
3,5,3',5'-, respectively. The authors were not able to relate the
difference in rates of elimination to chlorine substitution patterns.
In a study of the influence of molecular structure on the excretion
of 14 PCB congeners in mice, Gage and Holm (1976) found that the 4,4'-;
3,3',4',6'-; 2,3,2'.4',6'-; and 2,3,4,2',4',5'-isomers were eliminated
most rapidly. These compounds had at least one pair of ortho-meta
vicinal carbon atoms unsubstituted, a configuration thought to be
important for rapid metabolism and excretion. The most slowly eliminated
compounds were 2,4,5,2',4',5'- and 2,3,4,2',4'.5'-hexa-isomers.
Felt et al. (1977) examined the elimination of [14C]-2,5,4'-tri-CB
in rhesus monkeys. The monkeys were fed 550 mg of the compound in fruit
daily for 84 days. On the basis of total excreted and recovered
radioactivity, the half-life of 2,5,4'-tri-CB was found to be 4.5 to
4.8 days.
Bleavins et al. (1984) examined the excretion of PCBs in female
European ferrets given a single dose of 0.05 mg [i^C]-labeled Aroclor
1254 in food. The results showed that urinary excretion accounted for
Sl/10 of the quantity of PCB that was eliminated in the feces. Excretion
of PCBs was highest during the first week following dosing, when 22.1
and 1.8% of the absorbed dose was excreted in the feces and urine,
respectively.
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52 Section 4
4.2.4.3 Dermal
Data concerning the excretion of PCBs by humans or animals
following dermal exposure were not located.
4.2.4.4 Parenteral routes
Human. No data were located in the available literature.
Animal. Injection studies indicate that PCBs can be excreted
unmetabolized into the gastrointestinal tract. Yoshimura and Yamamoto
(1975) recovered unmodified tetra-CB from the duodenal contents of rats
injected intravenously with tetra-CB. Daily excretion for 4 days ranged
from 0.5 to 0.8% of the total dose per day. Goto et al. (1974) found
that 4.7 to 23.2% of injected PCBs were excreted unchanged into the
gastrointestinal tract by 10 days postdosing, with the excretion of a
penta-isomer greater than the excretion of di-, tri-, or tetra-isomers.
4.3 TOZICITT
Evaluation of the toxicity of Aroclors and other commercial PCB
mixtures is complicated by numerous factors, including differences in
isomer/congener/mixture composition and toxicity, differences in species
susceptibility, quantitatively inconsistent data, and varying degree of
contamination with toxic chemicals such as chlorinated dibenzofurans. In
addition, there is a lack or paucity of toxicological data for some of
the Aroclors (most of the studies were conducted with the higher
chlorinated Aroclors), and a paucity of data for the most sensitive
species (monkey and mink). Also, it should be recognized that PCBs to
which people may be exposed may be very different from the original PCB
mixture because of changes in congener and impurity composition
resulting from environmental and/or biological transformation. Because
of the aforementioned concerns, current data are considered inadequate
to differentiate between the toxicity and carcinogenicity of PCB
mixtures with any reasonable degree of confidence. Therefore, it is
assumed, for the purpose of health effects evaluation, that effects
resulting from exposure to a specific Aroclor are representative of
effects that may be produced by the other Aroclors. In the following
sections, data delineating the threshold region of the most toxic
Aroclor for specific end points are presented. Although the relative
contribution of the inhalation and dermal routes in occupational
exposures is unknown, health effects data for exposed workers are
discussed in the inhalation subsections.
4.3.1 Lethality and Decreased Longevity
4.3.1.1 Inhalation
Human. Pertinent data were not located in the available
literature.
Animal. Inhalation LCSQs of Aroclor were not located in the
available literature. Rozanova (1943) reported that all four rats
exposed to Solvol (a European PCB mixture) at concentrations of 10 g/m3
for 3 h became comatose and died, while 11 similar exposures at 0.5 g/m3
resulted in only one death. Liver and renal damage was noted along with
-------
Toxicological Data 53
congestion in the heart and spleen. Insufficient detail was available to
determine how the atmosphere was generated or what methods were used to
verify the concentration. Treatment-related mortality was not observed
in groups of 9 to 10 rats, 6 to 10 mice, 3 to 4 rabbits, 4 to 6 guinea
pigs or 1 cat that were exposed 7 h/day, 5 days/week to vapor
concentrations of 8.6 mg/m3 (0.83 ppm) Aroclor 1242 for 24 days,
5.4 mg/m3 (0.41 ppm) Aroclor 1254 for 121 days. 6.83 mg/m3 (0.66 ppm)
Aroclor 1242 for 120 days, 1.5 mg/m3 (0.11 ppm) Aroclor 1254 for
213 days, or 1.9 mg/m3 (0.18 ppm) Aroclor 1242 for 214 days (Treon et
al. 1956). It was necessary to heat the Aroclors to 55 to 138°C to
attain the above concentrations, and 8.6 mg/m3 Aroclor 1242 was
"approaching saturation" concentration. These concentrations may be low
as the technique used to estimate them was invalidated. Possible
contamination by PCDF was not reported.
4.3.1.2 Oral
Human. Pertinent data were not located in the available
literature.
Animal. Acute oral LD50 values for the PCBs covered by this
profile (Aroclors 1254, 1221, 1260, 1232, 1242, and 1248) are presented
in Table 4.1. No values for Aroclor 1016 were found in the available
literature. The lowest oral LD50 in rats was 1.01 g/kg for Aroclor 1254
as reported by Garthoff et al. (1981). In mink, the lowest LD50 was
between 0.75 and 1.0 g/kg for Aroclor 1221 as reported by Aulerich and
Ringer (1977). As seen from the data of Grant and Phillips (1974) and
Linder et al. (1974), immature rats appear to be more sensitive than
adult rats. The full range of LD50 values for all PCBs is greater, with
the lowest value of 0.5 g/kg for hexachlorobiphenyl in guinea pigs
(McConnell and McKinney 1978) and the highest value of 11.3 g/kg
reported for Aroclor 1262 in the rat (Fishbein 1974).
In mice maintained on diets that provided 1,000 ppm Aroclor 1254
for 14 days, 3/5 died of unspecified causes by day 15 (Sanders et al.
1974). All mice treated at 4,000 ppm died within 7 days after the onset
of treatment. No deaths occurred in five mice that were similarly
treated with 250 ppm.
For intermediate-exposure durations, the LC50 for Aroclor 1254 fed
to mink in the diet for 28 days ranged from 79 to 84 ppm and 47 to
58 ppm after a 7-day withdrawal period (Hornshaw et al. 1986). In mink
fed Aroclor 1254 for 9 months, the LC50 was 6.65 ppm (Ringer et al.
1981). Death generally was due to nonspecific hemorrhagic lesions.
Groups of 24 male rats that were fed diets containing 0, 25, 50, or
100 ppm Aroclor 1254 for 104 to 105 weeks experienced dose-related
decreased survival (92, 83, 58, and 46%, respectively) (NCI 1978). The
cause of death was not specified, and there was no effect on survival in
similarly treated female rats. There was no attempt to identify or
quantitate impurities. Decreased survival is not a universal finding in
chronic PCB studies, as survival was unchanged or increased in rats
treated with 100 ppm of 60% chlorine PCB mixtures (Aroclor 1260 and
Clophen A-60) via diet (Norback and Weltman 1985, Schaeffer et al.
1984).
-------
Table 4.1. Acute oral LD^s of Aroclon
Aroclor
1254
1221
1260
1232
1242
1248
NR
Species/strain
Rat/Wistar
Rat/Sherman
Rat/Osborne-Mendel
Mink/pastel
Rat/NR
Rat/Sherman
Mink/pastel
Rat/Sherman
Rat/NR
Rat/Sprague-Dawley
Rat/NR
Mink/pastel
Rat/NR
= not reported.
Sex/age
M/30 days
F/30 days
M/60 days
F/60 days
M/ 120 days
F/ 120 days
M/ weanling
NRa/adult
M/adult
NR/NR
NR/NR
F/NR
NR/NR
NR/adult
M/weanling
NR/NR
M/adult
NR/NR
NR/NR
NR/NR
LDW
(g/kg)
1.3
1.4
1.4
1.4
2.0
2.5
1.29S
4-10
1.01 (single dose)
I.S3 (5 doses over 2tt weeks)
1.99(5 doses, 1 day/week)
4
3.98
4.0
>0.7Sto3
11
References
Grant and Phillips 1974
Under et al. 1974
Garthoff et al. 1981
Aulerich and Ringer
Fishbein 1974
Nelson et al. 1972
Aulerich and Ringer
Linder et al. 1974
Fishbein 1974
Bruckner et al. 1973
Fishbein 1974
Aulerich and Ringer
Fishbein 1974
1977
1977
1977
u»
A
O
ft
*-.
§
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Toxicological Data 55
4.3.1.3 Dermal
Human. Pertinent data were not located In the available
literature.
Animal. Median lethal doses for single application of Aroclors to
the skin of rabbits ranged from >1,269 mg/kg for Aroclors 1242 and 1248
in 50% corn oil to <3,169 mg/kg for undiluted Aroclor 1221 as reported
by Nelson et al. (1972) and summarized by Fishbein (1974) (Table 4.2).
4.3.2 Systemic/Target Organ Tozlcity
4.3.2.1 Liver
Inhalation, human. Epidemiological studies and clinical surveys
indicate that occupational exposure to Aroclors has produced increases
in serum liver-related enzymes, particularly GGTP and SCOT (Ouw et al.
1976; Alvares et al. 1977; Fischbein et al. 1979, 1985; Baker et al.
1980; Smith et al. 1981a,b,c; Brown and Jones 1981; Maroni et al. 1981a;
Fischbein 1985; Emmett 1985; Lawton et al. 1985; Drill et al. 1981;
Kreiss 1985; Guzelian 1985). These increases show generally inconsistent
patterns, may be nonspecific, may be within the normal population range,
and have not been shown to be associated with hepatic dysfunction.
Alvares et al. (1977) found that the mean half-life of antipyrine
disappearance from blood was significantly lower in five workers who
were exposed to Aroclor 1016 (10.8 h) than in controls (15.6 h).
Asymptomatic hepatomegaly was reported by Maroni et al. (1981a). The
subjects of the aforementioned studies were primarily involved in
electrical equipment (e.g., capacitors, transformers) manufacturing and
repair, and many had measurable and often high serum levels of PCBs.
Monitoring data were reported only in some of the studies and do
not adequately characterize exposure levels because of limitations and
dissimilarities in sampling methods, durations, and locations; changes
in workplace ventilation and Aroclor formulations during the exposure
period; wide ranges in concentrations within and between studies without
indications of average levels; emphasis on correlating effects with
serum PCB concentrations rather than air concentrations of PCBs; and
unknown contribution of dermal exposure to total exposure. It appears,
however, that air concentrations of Aroclors were often <1 mg/m-*.
Fischbein et al. (1979, 1985) reported that capacitor manufacturing
plant workers who were exposed to various Aroclors (primarily 1242 and
1254) experienced 8-h time-weighted average (TWA) concentrations ranging
from 0.007 to 11.0 mg/m3. Liver-related indices were evaluated in 280 of
the workers; approximately 40% of the workers had been employed for
>20 years. Of the workers with plasma levels of higher chlorinated PCBs
>75 ppb, 8.3% had abnormally high SCOT levels compared with 1.6% of the
workers with plasma levels of higher chlorinated PCBs £75 ppb. Of the
workers with plasma levels of lower chlorinated PCBs >200 ppb, 10.8% had
abnormally high SCOT levels compared with 1.2% of the workers with
plasma levels of lower chlorinated PCBs £200 ppb. The differences were
statistically significant, but remained significant only for females
when sexes were analyzed separately. An increased prevalence of abnormal
GGTP levels and weak but statistically significant correlations between
serum concentrations of PCBs and GGTP were also reported. Greater than
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56 Section 4
Table 4.2. Acute dermal LDM values of
Aroclors in rabbits
Aroclor
1221
1232
1242
1248
1260
Vehicle
Undiluted
Undiluted
Undiluted
Undiluted
50% corn oil
LD50
(mg/kg)
>2000 <3469
>1260 <2000
>794 <1269
>794 <1269
>1260 <2000
Source: Fishbein 1974.
-------
Toxicologies! Data 57
90% of the values for SCOT and other liver-associated enzymes were
within normal laboratory limits and the prevalence of abnormal values
was comparable to the general population. Limitations of this study,
including broad exposure categories, lack of an unexposed control group,
and lack of correction for confounding variables such as alcohol
consumption, indicate that the data, while suggestive, should not be
interpreted as demonstrating a relationship between SCOT levels and
plasma levels of PCBs.
Inhalation, animal. Reversible degenerative lesions of the liver
were observed in rats, mice, rabbits, cats, and guinea pigs exposed to
1.5 mg/m-3 (0.11 ppm) Aroclor 1254 vapor 7 h/day, 5 days/week for a
213-day period (Treon et al. 1956). Exposure to Aroclor 1242 for
7 h/day, 5 days/week at 1.9 mg/m3 (0.18 ppm) for 214 days or 8.6 mg/m3
(0.83 ppm) for 24 days did not produce histological effects in the liver
or other viscera. It was necessary to heat the Aroclors to attain the
concentrations used in this study.
Oral, human. Serum levels of PCBs and GGTP were positively
correlated in Triana, Alabama, residents (Kreiss et al. 1981).
Consumption of contaminated fish was the only known source of PCS
exposure. The population was also exposed to DDT via consumption of fish
and the strongest correlations were between serum levels of PCBs and
DDT, but the effects of DDT residues on the metabolism or toxicity of
PCBs are unknown.
Oral, animal. Carter (1985) exposed groups of 12 male weanling
Charles River rats to 0, 4, 8 or 16 ppm of Aroclor 1254 in the diet for
4 days. Relative liver weights were significantly increased at >8 ppm,
and serum levels of HDL cholesterol were significantly increased at
16 ppm. Histological examinations were not performed.
Litterst et al. (1972) exposed groups of six male Osborne-Mendel
rats to Aroclors 1260, 1254, 1248, or 1242 in the diet at concentrations
of 0, 0.5, 5.0, or 500 ppm for 4 weeks. Increased microsomal
nitroreductase and demethylase activities occurred at >0.5 ppm,
increased pentobarbital hydroxylation and relative liver weight occurred
at >50 ppm, and increased liver triglycerides occurred at 500 ppm.
Dietary exposure to 5 or 25 ppm Aroclor 1242 for 2, 4, or 6 months
produced increased hepatic microsomal hydroxylase activity and
histochemically discernible lipid content of hepatocytes in groups of
six male Sprague-Dawley rats (Bruckner et al. 1974). Increased relative
liver weight was observed at 25 ppm at 4 and 6 months and at 5 ppm at
4 months.
Frank histological effects in the liver (e.g., fatty degeneration)
occurred in rats exposed to >20 ppm Aroclor 1254 or 1260 for 28 days
(Chu et al. 1977), rats exposed to £20 ppm Aroclor 1254 or 1260 for
8 months (Kimbrough et al. 1972), and mice exposed to 37.5 ppm but not
3.75 ppm Aroclor 1254 for 6 months (Roller 1977).
In a study in which 4 male and 18 female rhesus monkeys were fed
diets containing Aroclor 1248, Barsotti et al. (1976) conducted
autopsies on one female monkey that died after being fed 2.5 ppm of
Aroclor 1248 3 days and on one female monkey that died after being
fed 5.0 ppm of Aroclor 1248 for 310 days. Hepatic effects in both
-------
58 Section 4
monkeys included focal areas of necrosis, enlarged hepatocytes, and
lipid droplets. Although only one animal per dose was examined, these
effects must be regarded as treatment-related because of the
characteristic nature of the hepatic response. Also, similar effects on
the liver were observed in an earlier study by Allen (1975) in which the
animals received Aroclor 1248 in the diet at levels of 100 and 300 ppm
for 2 or 3 months.
Chronic dietary studies were conducted with rats exposed to 25 to
100 ppm Aroclor 1254 for 2 years (NCI 1978, Morgan et al. 1981, Ward
1985), 100 ppm Aroclor 1260 for 16 months followed by 50 ppm for
8 months, and then no treatment for 5 months (Norback and Weltman 1985)
or 100 ppm Aroclor 1260 for 21 months (Kimbrough et al. 1975).
Treatment-related nonproliferative liver lesions or nonproliferative
liver lesions that did not progress to neoplasms after 1 year were not
described in these studies.
The effects of chlorination and chemical composition of PCBs with
regard to the dose effects relation of liver toxicity after subchronic
exposure are indicated by the data of Biocca et al. (1981). In this
study, hepatotoxic effects were observed in mice after 5 weeks of
maintenance on diets containing 0.3 ppm of 3,4,5-symmetrical
hexachlorobiphenyl, while similar effects were observed only after
30 ppm of 2,4,5-symmetrical hexachlorobiphenyl and 100 ppm of 2,4,6-
symmetrical hexachlorobiphenyl, and no effects were noted after 300 ppm
of 2,3,6-symmetrical hexachlorobiphenyl. Similar dependence of liver
toxicity on the chemical composition of the PCB mixture would be
anticipated following chronic exposure in mice and other species.
None of the above studies reported possible contamination of the
Aroclor with PCDF.
Dermal, human. A study of capacitor workers, already discussed
under inhalation (Sect. 4.3.2.1: Systemic/Target Organ Toxicity, Liver,
Inhalation, human), provided PCB exposure measurements of 48-275 Mg/m3
in workroom air and 2-28 /ig/cm2 of skin surface on the palms of the
workers' hands (Maroni et al. 1981a,b). The authors concluded that much
of the absorption of PCBs occurred through the skin. Of the 80 exposed
workers, 16 had some evidence of liver involvement including
asymptomatic hepatomegaly and/or elevated (to slightly above normal
range) serum levels of GGPT, SCOT, or SGPT. No control group was
included in the study. The findings were considered by the authors to be
indicative of hepatic microsomal induction. Drill et al. (1981)
concluded that the serum enzyme levels reflected random variations from
normal, but did not discuss the finding of hepatomegaly.
Dermal, animal. Aroclor 1260 in isopropanol vehicle was applied to
the shaved backs of groups of four female New Zealand rabbits daily
5 days/week at a dose of 118 mg/day for 38 days (Vos and Beems 1971) or
120 mg/day for 28 days (Vos and Notenboom-Ram 1972). Histological
alterations were produced in the livers, including centrolobular
degeneration and liver cell atrophy, focal hyalin degeneration of the
cytoplasm of the hepatocyte, enlarged nuclei, and loss of glycogen.
Aroclor 1260 used in these experiments was reported to be free of PCDF
contamination.
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Toxicological Data 59
General discussion. The liver is the organ most often implicated
in the toxicity of Aroclors in animals. Hepatic effects have been
observed in numerous studies with exposed rats, mice, guinea pigs,
rabbits, dogs, and monkeys, but rats have been tested most extensively.
The effects appear to be reversible at low doses (Treon et al. 1956),
are similar among species, and include hepatic microsomal enzyme
induction, increased serum levels of liver-associated enzymes indicative
of possible hepatocellular damage, liver enlargement, fat deposition,
and necrosis. Microsomal enzyme induction is the most sensitive
indicator of hepatic alterations, but this effect is not necessarily
adverse and few studies were designed to define minimum effective doses
of Aroclors. The liver enlargement is associated with hepatocyte
enlargement and an increase in smooth endoplasmic reticulum and/or
increased enzymatic activity. Proliferative lesions in the liver have
been attributed to Aroclor treatment (Sect. 4.3.6 on carcinogenicity in
this section). The hepatic effects of Aroclors in animals appear to be
typical of chlorinated hydrocarbons.
Histologically documented liver damage is a consistent finding
among PCB-exposed animals. Studies of Aroclor-exposed workers provide
inconsistent but suggestive evidence for subclinical increases in serum
enzymes that are indicators of possible liver microsomal enzyme
induction or possible hepatocellular damage (e.g., GGPT, SCOT) (EPA
1988a, Kreiss 1985, Drill et al. 1981). Hepatic dysfunction has not been
demonstrated in PCB-exposed workers. That hepatic alterations have been
inconsistently observed in humans may be related to the fact that many
of the studies (particularly the earlier ones) did not account for
confounding variables, such as alcohol consumption, exposure to
additional chemicals, or previous medical histories, or may be an
artifact of the relative insensitivity of the standard biochemical tests
of liver damage (e.g., SCOT) as compared with biopsy evaluation (Letz
1983, Drill et al. 1981). Drill et al. (1981) concluded that SCOT and/or
GGPT appear to be the most sensitive indicators of PCB exposure in
humans, and that changes in liver enzymes may occur at levels below
those at which chloracne occur. Abnormal liver function and some
hepatomegaly have been documented in Yusho and Yu Cheng patients, but
PCDFs, polychlorinated quaterphenyls, and perhaps other contaminants
(e.g., chlorinated diphenyl ethers) are significant etiologic factors
(Fischbein 1985).
Aroclors are commonly used to induce hepatic enzymes in animal
studies with other chemicals. Exposures in these studies are not
representative of realistic human exposures, as large doses are usually
given by intraperitoneal injection or gavage to obtain maximal enzyme
induction. Induction of enzymes by PCBs occurs in both the cytochrome
P-450 and P-448 systems, has been observed in humans, and is not
restricted to the liver (Letz 1983). Implications of enzyme induction
for human health include the possibility of disease secondary to the
increased metabolism of endogenous substances (such as hormones) and
increased metabolic activation of exogenous substances, possible
protective effects secondary to the increased metabolic detoxification
of exogenous substances, and the interference with medical therapy due
to increased metabolism of administered drugs (Letz 1983).
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60 Section 4
Safe et al. (1985a) reviewed data concerning the mechanism of PCB
induction of liver microsomal enzymes. The activity of individual PCBs
depends on their structure. The most active congeners are those
substituted at both para and at two or more meta positions and include
3,4,4',5-tetra-, 3,3',4,4'-tetra-, 3,3',4,4',5-penta-, and
3,3',4,4',5,5'-hexachlorobiphenyl. The coplanar PCBs induce rat liver
microsomal aryl hydrocarbon hydroxylase and cytochromes P-450a, P-450c,
and P-4SOd, thus resembling 3-methylcholanthrene and 2,3,7,8-TCDD in
their mode of microsomal enzyme induction. Mono-ortho- and di-
orthochloro analogs of coplanar PCBs exhibit a mixed type of enzyme
induction similar to Aroclor 1254. These PCBs induce aryl hydrocarbon
hydroxylase, dimethylaminoantipyrine, N-demethylase, and cytochromes
P-450a through P-450e. Results of quantitative structure-activity
relationships showed a correlation between aryl hydrocarbon hydroxylase
induction activity and binding affinity for the 2,3,7,8-TCDD cytosolic
receptor protein, with the order of activity as follows: coplanar PCBs >
3,4,4',5-tetrachlorobiphenyl « mono-ortho coplanar PCBs > di-ortho
coplanar PCBs. Support for the receptor-mediated mechanism of action was
found when the coplanar and mono-ortho coplanar PCBs were administered
to C57BL/6J and DBA/2J mice. CS7BL/6J mice contain much higher
concentrations of the Ah receptor than do DBA/2J mice. The PCBs induced
aryl hydrocarbon hydroxylase in the responsive C57BL/6J mice but not in
the unresponsive DBA/2J mice. Although there is general agreement
regarding the role of the Ah receptor in microsomal enzyme induction,
the role of Ah receptor binding in the toxicity of PCBs and other
halogenated aromatic hydrocarbons is unclear.
4.3.2.2 Cutaneous tissues
Inhalation, human. Effects such as chloracne, skin rashes, and
burning eyes and skin have been associated with occupational exposure to
Aroclors (Meigs et al. 1954; Ouw et al. 1976; Fischbein et al. 1979,
1982, 1985; Baker et al. 1980; Smith et al. 1981a.b,c; NIOSH 1977a; EPA
1988a; Drill et al. 1981; Kimbrough 1987a). Monitoring data do not
adequately characterize exposure levels for the reasons indicated in
Sect. 4.3.2.1 on liver effects in humans after inhalation exposure.
Correlations between chloracne and duration of exposure or blood
concentrations of PCBs are poor or nonexistent, and the actual incidence
of chloracne is unknown but appears to be low. Drill et al. (1981)
concluded that individuals with blood PCB levels >200 ppb have an
increased risk of chloracne and that chloracne may occur more frequently
among workers exposed to PCBs that have been heated and to PCBs that
have >54% chlorination. The available evidence, however, cannot be used
to conclude that 200 ppb represents a threshold for chloracne. The
conclusions of Drill et al. (1981) are based on Kanechlor as well as
Aroclor toxicity data. As chloracne is reported frequently among workers
who were exposed to Kanechlors, the higher chloracnegenic potential of
Kanechlors and heated Aroclors may be related to higher levels of PCDFs
and polychlorinated quaterphenyl contaminants (Drill et al. 1981).
Fischbein et al. (1979, 1982) conducted a clinical survey of 289
capacitor manufacturing workers (153 male, 136 female) who were exposed
to 0.007-11 mg/m3 concentrations of various Aroclors; 20% of the worke*
had been employed for 5-10 years and 39% for >20 years. Sixty-nine (45
-------
Toxicological Data 61
male and 75 (55%) female workers had a history of dermatological
complaints. Physical examination revealed that 59 males (39%) and 48
females (35%) had abnormal dermatological findings. The most prevalent
skin abnormalities were erythema, dryness, thickening, and eye
abnormalities (conjunctival redness, palpebral hyperpigmentation, and
edema); nonadolescent acneform eruptions were observed in 16 individuals
[7 males (5%) and 9 females (7%)]. A subgroup of 42 workers (22 males,
20 females) with skin effects that clinically were thought to be related
to PCB exposure in particular (e.g., hyperpigmentation, comedones,
chloracne) was compared with an unspecified population; a difference was
found between the mean plasma concentrations of higher chlorinated PCBs
in males with and without skin abnormalities. The difference was
statistically significant using a Student's t-test adjusting for unequal
variances (P - 0.03), but not using the t-test to compare mean log
plasma concentrations (P - 0.07) or using nonparametric tests. These
data suggest an association between dermatologic effects and plasma
levels of higher chlorinated PCBs.
Thirty-four workers who were exposed to Aroclor 1242 at
concentration between 0.32 and 2.22 mg/m^ for 5 to 23 years in an
electrical plant complained of burning of the eyes, face, and skin; five
had eczematous rashes on the hands and legs (Ouw et al. 1976). The
Aroclor 1242 was reported to be free of impurities.
Inhalation, animal. Pertinent data were not located in the
available literature.
Oral, human. Pertinent data were not located in the available
literature.
Oral, animal. Cutaneous effects occurred in rhesus monkeys fed
diets that contained Aroclors for subchronic durations (Allen and
Norback 1973, Allen et al. 1974a, Allen 1975, Barsotti and Allen 1975,
Barsotti et al. 1976, Thomas and Hinsdill 1978, Becker et al. 1979,
Allen et al. 1979, McNulty et al. 1980). These include facial
(particularly periorbital) edema, purulent discharge from the eyes,
chloracne, and alopecia. The effects appear to be reversible and have
been produced by diet exposures as low as 2.5 ppm Aroclor 1248 for 1 to
6 months (Barsotti and Allen 1975) and 3 ppm Aroclor 1242 for 6 months
(Becker et al. 1979). NOAELs were not identified in the available
studies.
In the Barsotti and Allen (1975) study, rhesus monkeys were fed
diets containing 2.5 or 5.0 ppm Aroclor 1248 for 1 year. The animals
exposed to 2.5 ppm (all females) developed periorbital edema, alopecia,
erythema, and acneform lesions of the face and neck within 1 to 2
months. The males treated at 5.0 ppm had only moderate periorbital edema
and erythema.
Thomas and Hinsdill (1978) fed 0, 2.5, and 5.0 ppm Aroclor 1248 to
adult female rhesus monkeys. All eight monkeys in each Aroclor-treated
group developed alopecia, chloracne, and facial edema after 6 months of
treatment.
In the Becker et al. (1979) study, six young (7 to 8 months old)
monkeys were fed diets containing 0, 3, 10, 30, or 100 ppm Aroclor 1242
(two were fed 10 ppm). Facial changes (palpebral swelling and erythema
-------
62 Section 4
but no loss of hair) were evident by the end of the second month at
>10 ppm and in the sixth month at 3 ppm; mortality was 4/6 by day 245,
including the monkey fed 3 ppm.
Rats exposed to Aroclor 1254 in the diet developed alopecia, facial
edema, and exophthalmos after 104 weeks of 50 ppm and 72 weeks of 50 ppm
(NCI 1978); these effects did not occur after 104 weeks of 25 ppm.
In a single-dose study, thickening and erythema of the pinna of the
ear occurred in mice exposed to 200 ppm of Aroclor 1254 in the diet for
23 weeks (Bell 1983).
All of the above studies did not report possible impurities.
Dermal, human. Pertinent data were not located in the available
literature.
Dermal, animal. Daily application of 118 mg Aroclor 1260 (free of
PCDF) in isopropanol vehicle to the shaved backs of four female New
Zealand rabbits 5 days/week for 38 days produced thickening of the skin
and acneform lesions resulting from hyperplasia and hyperkeratosis of
the epidermal and follicular epithelium (Vos and Beems 1971). These
results were verified in another similarly designed study (Vos and
Notenboom-Ram 1972).
General discussion. Relatively small groups of animals were tested
in most of the studies, but the cutaneous effects are well
characterized. The cutaneous effects in occupationally exposed humans
are generally consistent with the animal data, but effect levels cannot
be ascertained and the contribution of direct skin exposure or
contaminants cannot be evaluated with the information reported in the
papers.
4.3.2.3 Immunological effects
Inhalation, human. Significant alterations in various globulin
fractions have not been observed in Aroclor-exposed workers (Ouw et al.
1976; Smith et al. 1981a,b,c). No difference in the incidence of
positive responses was found during skin hypersensitivity testing with
mumps and trichophyton in Aroclor-exposed switchgear workers and
unexposed workers (Mosley and Emmett 1984). Elevations in total white
blood cells associated with decreased polymorphonuclear cells and
increased lymphocytes, monocytes, and eosinophils were measured in
capacitor workers 1 year before discontinuance of Aroclor use in the
operation (Lawton et al. 1985). The findings were difficult to interpret
because they were also associated with dichlorodiphenyldichloroethylene
(DDE) exposure.
Inhalation, animal. Pertinent data were not located in the
available literature.
Oral, human. Pertinent data were not located in the available
literature.
Oral, animal. Female guinea pigs maintained on diets that
contained 50 ppm of Aroclor 1260 for 6 weeks had significantly lowered
tetanus autotoxin titers, circulating leukocytes and lymphocytes, and
-------
Toxicologies! Data 63
thymus atrophy (Vos and van Genderen 1973). Exposure to 10 ppm Aroclor
1260 in the diet for 8 weeks produced splenic atrophy in guinea pigs
(Vos and de Roij 1972). NOAELs were not identified in these studies. The
Aroclor 1260 used in these studies was reported to be free from PCDF
impurities.
Thomas and Hinsdill (1978) exposed groups of 5 to 8 female rhesus
monkeys to 0, 2.5, or 5.0 ppm Aroclor 1248 in the diet for 11 months.
Significantly lower antibody response to sheep red blood cells occurred
at- 5.0 ppm. There was no treatment-related effect on antibody response
to tetanus toxoid.
Barsotti et al. (1976) also found evidence of an immunological
effect in rhesus monkeys fed 2.5 or 5.0 ppm Aroclor 1248 in the diet for
7 months prior to mating and during pregnancy. Monkeys developed
shigellosis during and after treatment, indicating an increased
susceptibility to infection.
Thomas and Hinsdill (1978) also fed Aroclor 1248 to mice at 100 or
1,000 ppm in the diet for 3 to 5 weeks. The mice had enhanced
sensitivity to Salmonella typhimurium and endotoxin, indicating lowered
resistance to infection.
Dermal, human. Pertinent data were not located in the available
literature.
Dermal, animal. Dermal application of 120 mg/day Aroclor 1260
(free of PCDF impurities) in isopropanol 5 days/week for 4 weeks
produced moderate thymic atrophy in rabbits (Vos and Notenboom-Ram
1972). Similar application of 118 mg/day Aroclor 1260 for 38 days
produced histological atrophy of the thymus cortex and a reduction in
the number of germinal centers in the spleen and lymph nodes in rabbits
(Vos and Beems 1971).
General discussion. Immunotoxic effects of PCBs in humans have not
been clearly demonstrated. Studies in animals, however, have shown
effects on the immune system. Immunosuppression was observed in monkeys
that received Aroclor 1248 in the diet at concentrations as low as
5.0 ppm (Thomas and Hinsdill 1978). Treatment of rodents with oral or
dermal doses of Aroclors, non-Aroclor PCBs, and/or individual PCB
congeners that have a different composition than those covered by this
profile has also produced effects on the immune system. This is
illustrated in the study by Biocca et al. (1981) in which a decrease in
thymus weight was observed in mice exposed to 3,4,5-symmetrical
hexachlorobiphenyl for 5 weeks in the diet at 10 ppm, compared with
similar effects produced at levels of 300 ppm for 2,4,5- or 2,4,6-
symmetrical hexachlorobiphenyl or at 167 ppm Aroclor 1242 in the diet of
mice in a 6-week study (Loose et al. 1978a,b). These effects include
immunosuppression as measured by increased mortality to Salmonella
typhosa lendotoxin and Plasoodium berghei in mice given 167 ppm Aroclor
1016 or 1242 in the diet for 6 weeks (Loose et al. 1978a,b), and
increased mortality caused by S. typhiourium endotoxin in mice that were
given 100 or 1,000 ppm Aroclor 1248 in the diet for 5 weeks (Thomas and
Hinsdill 1978). PCBs also caused splenic, thymic, and lymph node atrophy
in rats (Allen et al. 1975, Allen and Abrahamsom 1973, Parkinson et al.
1983).
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64 Section 6
Although PCBs appear to be immunosuppressive In animals, the effec
of PCBs on immune system function in humans has not been adequately
evaluated. Based on animal splenic and lymphoid system histological
alterations, Drill et al. (1981) speculated that significant
immunosuppression in humans may occur only at high dosages secondary to
malnutrition (i.e., via general toxic responses such as decreased food
intake, decreased body weight, or decreased body weight gain). From
their results in monkeys, Thomas and Hinsdill (1978) concluded that the
occasional ingestion of food contaminated with 5 ppm PCBs by humans
would probably not result in immunosuppressive effects measured by
decreased antibody titers.
Immunotoxicity of PCBs appears to be dependent upon expression of
the aromatic hydrocarbon receptor and on the ability of PCBs to bind to
the receptor (EPA 1988a). The receptor binding affinity of PCBs is
dependent on the molecular conformation that is determined by the
chlorine substitution pattern.
4.3.2.4 Thyroid
Inhalation. Pertinent data were not located in the available
literature.
Oral, human. Pertinent data were not located in the available
literature.
Oral, animal. Rats exposed to Aroclor 1254 for 4 to 12 weeks
experienced thyroid alterations that included enlargement, reduced
follicular size, follicular cell hyperplasia, and accumulation of
colloid droplets and large, abnormally shaped lysosomes in the
follicular cells (Collins et al. 1977; Collins and Capen 1980b,c; Kasza
et al. 1978). The thyroid alterations resulted in reduced serum
thyroxine levels and appear to be reversible after cessation of
exposure. None of these studies reported the purity of the Aroclor 1254
sample used.
Collins and Capen (1980b) exposed groups of six male Osbome-Mendel
rats to 0, 5, 50, or 500 ppm Aroclor 1254 in the diet for 4 weeks.
Histological and ultrastruetural effects consistent with those described
above occurred at >5 ppm, and reduced serum thyroxin occurred at
>50 ppm. A NOAEL for thyroid alterations cannot be discerned from the
available studies.
Dermal. Pertinent data were not located in the available
literature.
General discussion. Although effects of Aroclor exposure on the
thyroid have been investigated in only a few studies, this gland is an
unequivocal target of Aroclor in rats. The lowering of serum thyroxine
by Aroclors appears to be the combined result of a direct effect on
thyroid follicular cells with an interference in hormone secretion plus
an enhanced peripheral metabolism of thyroxine (Collins et al. 1977).
Ultrastruetural lesions in thyroid follicular cells and reductions
in serum levels of thyroid hormones (thyroxine and triiodothyronine)
occurred in neonatal and weanling rats whose dams were fed diets
containing 50 or 500 ppm Aroclor 1254 throughout gestation and lactatior
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Toxicological Data 65
(Collins and Capen 1980a). These authors also reported that other
studies have found that decreased reproductive performance and
interference in growth and development occurred in man and animals that
were rendered hypothyroid and that PCBs enhance the peripheral
metabolism and excretion of thyroxine-glucuronide in the bile. These
findings and the thyroid effects in Aroclor-exposed adult rats
summarized previously suggested to Collins and Capen (1980a) that some
of the we11-documented PCB-related disturbances in reproduction, growth,
and development may be related to alterations in thyroid structure and
function in the dam, fetus, or neonate.
4.3.2.5 Stomach
Effects on the stomach have been studied only in animals exposed
orally. Oral administration of Aroclor 1248 (Allen and Norback 1973;
Allen et al. 1974a,b; Allen 1975; Barsotti and Allen 1975) and Aroclor
1242 (Becker et al. 1979) to monkeys produced gastritis, which
progressed to hypertrophy and hyperplasia of the gastric mucosa. Related
effects include mucous-filled cysts that penetrate the muscularis
mucosa. These effects were initiated by exposures as low and/or short as
a single gavage dose of 1.5 g/kg of Aroclor 1248 (Allen et al. 1974a),
25 ppm of Aroclor 1248 in the diet for up to 1 year (Barsotti and Allen
1975), and 3 ppm of Aroclor 1242 for 71 days (Becker et al. 1979).
The Aroclor-induced gastric lesions occurred only along the greater
curvature of the stomach (not in the cardiac or pyloric regions, which
are more usual regions for gastric effects), did not occur in other
sections of the gastrointestinal tract, and have not been observed in
species other than monkeys (Becker et al. 1979, Drill et al. 1981).
These gastric effects may therefore be species specific. Aroclor 1254-
induced metaplasia and adenocarcinoma in the glandular stomach of F344
rats have been reported (Morgan et al. 1981) (Sect. 4.3.6 on
carcinogenicity in this section). These studies did not report the
purity of the Aroclor sample used.
4.3.2.6 Porphyria
Inhalation, human. Exposure-related urinary porphyrin excretion,
porphyrin-related disease, or cases of porphyria cutaneous tarda have
not been reported in several clinical studies of Aroclor-exposed workers
(Alvares and Kappas 1979; Fischbein et al. 1979; Smith et al.
1981a,b,c). A clinical study by Colombi et al. (1982), however, reported
a marked increase in the excretion of urinary porphyrins by Aroclor-
exposed workers whose blood levels of PCBs were at least ten times
higher than expected in a population without occupational exposure to
PCBs. The relative proportions of the urinary porphyrins did not differ
from those in controls, indicating that the increase was due to a
generalized increase in porphyrin synthesis by the liver, probably
because of induction of liver microsomal enzymes. No evidence of
porphyria was seen in these workers, but the investigators pointed out
that a similar increase in urinary porphyrin excretion In experimental
animals is followed by porphyria if administration of Aroclors
continues.
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66 Section 4
Inhalation, animal. Pertinent data were not located in the
available literature.
Oral, human. Pertinent data were not located in the available
literature.
Oral, animal. Groups of six male Sprague-Dawley rats were treated
with 0, 5, or 25 ppm of Aroclor 1242 (purity not reported) in the diet
for 2, 4, or 6 months (Bruckner et al. 1974). Urinary coproporphyrin
levels were increased in rats treated at both concentrations.
In rats fed 100 ppm Aroclor 1254 in the diet, Goldstein et al.
(1974) found that liver microsomal P-450 concentrations and liver weight
were increased maximally by 1 week, but that the onset of porphyria and
induction of ALA synthetase was delayed until 2-7 months of treatment. A
marked accumulation of uroporphyrins occurred in the liver, and urinary
excretion of coproporphyrin and other porphyrins was increased, with the
largest increase in uroporphyrins. The uroporphyrins in liver and urine
of the treated rats consisted primarily of 8-carboxy- and 7-
carboxyporphyrins. The disproportionate increase in hepatic and urinary
uroporphyrins could have been due, in part, to a decrease in
uroporphyrinogen decarboxylase activity (Goldstein et al. 1974, Hill
1985).
Dermal, human. Pertinent data were not located in the available
literature.
Dermal, animal. Fecal coproporphyrin was elevated in female New
Zealand rabbits that received a 120-mg application of Aroclor 1260 to
shaved backs 5 days/week for 4 weeks (Vos and Notenboom-Ram 1972). Fecal
coproporphyrin and protoporphyrin were increased in rabbits similarly
treated with 118 mg/day Aroclor 1260 5 days/week for 36 days (Vos and
Beems 1971). The Aroclor 1260 used in these studies was free of PCDF.
General discussion. Induction of ALA synthetase (a rate-limiting
enzyme in heme synthesis) and inhibition of uroporphyrinogen
decarboxylase are the mechanisms of porphyrogenic action of other
polyhalogenated aryl hydrocarbons (Colombi et al. 1982, Drill et al.
1981, Hill 1985). It has been suggested that the changes in porphyrin
metabolism are triggered by the induction of liver microsomal enzymes
(Colombi et al. 1982). The results of Goldstein et al. (1974) in rats
fed Aroclor 1254 in the diet suggest that Aroclors may produce porphyria
in a similar manner. Although porphyria has not been reported in
Aroclor-exposed humans, increased urinary excretion of porphyrins has
been observed in one study of occupationally-exposed humans (Colombi et
al. 1982), and evidence of induction of hepatic microsomal enzymes has
also been observed (Sect. 4.3.2.1 on Systemic/Target Organ Toxicity,
Liver). There are no data to indicate that a progression from these
alterations to porphyria would occur as a consequence of continued
occupational exposure to Aroclors, but such a progression has been
demonstrated in orally exposed animals (Goldstein et al. 1974). Drill et
al. (1981) raised the possibility that PCBs, via induction of ALA
synthetase, might be capable of precipitating an attack of porphyria in
patients suffering from acute, intermittent porphyria. Chronic hepatic
porphyria and porphyria cutanea tarda are associated with exposure to
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Toxicological Data 67
other polyhalogenated aryl hydrocarbons, including polybrominated
biphenyls and 2,3,7,8-TCDD (Hill 1985).
4.3.2.7 Kidney
The only study that reported effects on the kidneys was Vos and
Beems (1971). In this study, Aroclor 1260 in isopropanol vehicle was
applied to the shaved backs of New Zealand rabbits for 5 days/week at a
dose of 118 mg/day for 38 days. Hydropic degeneration of the convoluted
tubules, destruction of tubular epithelial cells, tubular dilation, and
proteinaceous casts were observed. No mention of kidney effects was made
in the study by Vos and Notenboom-Ram (1972), in which Aroclor 1260 was
applied to the shaved backs of rabbits at 120 mg/day, 5 days/week for
28 days.
4.3.3 Developmental Toxicity
4.3.3.1 Inhalation
Human. Fifty-one infants born to women employed at two capacitor-
manufacturing facilities with a history of high exposure to Aroclors
1254, 1242, and/or 1016 had mean birth weights and mean gestational ages
that were lower than infants born to women who had worked in low-
exposure areas (Taylor et al. 1984). The differences were small (153 g
and 6.6 days), and the birth weight difference appears to have resulted
from the shortened gestation rather than from a retardation of
intrauterine growth. The high-exposure workers were directly exposed to
Aroclors during the manufacturing process for at least 1 year prior to
birth of the infant; the workers with low exposure were employed in
areas where Aroclors were not used directly. The results of this study
are considered to be suggestive but inconclusive because the effects
were small and confounding factors such as smoking and alcohol
consumption, prenatal care, underlying medical conditions, maternal
height, and previous history of low birth weight were not considered.
Animal. Pertinent data were not located in the available
literature.
4.3.3.2 Oral
Human. Birth weight, length, head circumference, gestational age,
and neonatal behavior were evaluated in 313 newborn infants (Fein 1984,
Fein et al. 1984, Jacobson et al. 1984a). Of these infants, 242 were
born to mothers who had consumed moderate to large quantities of Lake
Michigan fish sometime during their lives, and 71 were born to mothers
who did not consume Lake Michigan fish. Mean (± standard deviation) fish
consumption and duration of consumption were 6.7 ± 5.8 kg/year and
15.9 ± 9.1 years, respectively; consumption during pregnancy was
4.1 ± 4.4 kg/year. Maternal serum PCB concentrations averaged 5.5 ± 3.7
ng/mL, which reportedly is comparable to those for other midwestern area
samples, and umbilical cord serum PCB levels averaged 2.5 ± 1.9 ng/L.
Both maternal consumption of fish and levels of PCBs in cord serum were
positively correlated with lower birth weight, smaller head
circumference, and shorter gestation (Fein et al. 1984). Infants of
-------
68 Section 4
mothers who had consumed contaminated fish were, on the average, 190 g
lighter, had head circumferences 0.6 cm less, and were born 4.9 days
earlier than infants of mothers who had not consumed contaminated fish.
Similar values were determined when infants with cord serum levels
>3 ng/mL were compared with infants whose cord levels were <3 ng/mL (the
analytical quantification limit) (160 g lighter, 0.6 cm less in head
circumference, 8.8 days less in gestational age). Head circumference was
significantly smaller in both analyses even after birth weight and
gestational age were statistically controlled. Contaminated fish
consumption was also positively correlated with impaired autonomic
maturity, increased numbers of abnormal reflexes, and decreased range of
state (Jacobson et al. 1984b). Range of state is a neurological category
that includes peak of excitement, rapidity of buildup, irritability, and
lability of state.
Jacobson et al. (1985) studied the effect of intrauterine exposure
or exposure through breast milk to PCBs on visual recognition memory and
preference for novelty in 123 infants. Measures of exposure included
reports by mothers of contaminated fish consumption and analysis of cord
serum levels and breast milk levels of PCBs. Reports of fish consumption
and cord serum levels were predictors of poor visual recognition memory,
while breast milk levels were not. There was a dose-related decrease in
fixation to novelty: cord serum levels of 0.2 to 1.1 ng/mL were
associated with mean scores of 61%, 1.2 to 2.2 ng/mL with mean scores of
60%, 2.3 to 3.5 ng/mL with scores of 57%, and 3.6 to 7.9 ng/mL with
scores of 50%.
Limitations of these studies include lack of analysis for chemicals
other than PCBs, failure to report maternal and cord serum PCB levels
based on fish consumption, correlation of effects with fish consumption
but not cord serum PCB levels, PCB blood levels within the range of the
general population, and/or unknown effects of maternal genetic makeup,
lifestyle, and acute illness.
Rogan et al. (1986) examined birth weight, head circumference, and
the results of behavioral tests in 930 children. At birth, samples of
placenta, maternal and cord serum, and milk were collected and analyzed
for PCBs. There was no correlation between birth weight or head
circumference with PCB levels. Levels of PCBs in milk fat at birth of
3.5 to >4 ppm, but not <3.49 ppm, were significantly correlated with
less muscle tone, decreased activity, and abnormal reflexes. The levels
of PCBs to which these infants were exposed were probably as high as
those encountered in the general population. Follow-up evaluation of the
same children showed no adverse effects on weight or frequency of
physician visits for various illnesses (Rogan et al. 1987). Because of
confounding exposure to DDE, the effects on neonatal behavior cannot be
attributed solely to PCBs.
Although these studies have several limitations, they provide
strongly suggestive but not yet conclusive evidence of behavioral
effects of PCBs in humans.
Animal. Rabbits were exposed to 0, 1.0, or 10.0 mg/kg/day and
12.5, 25.0, or 50 mg/kg/day Aroclor 1254 (purity not reported) by gavage
on days 1 to 28 of pregnancy in separate experiments (Villeneuve et al.
1971). Abortions, stillbirths, and maternal deaths occurred at
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Toxicologies! Data 69
>12.5 mg/kg/day, but there were no treatment-related teratogenic effects
at any dose level. It was noted that unpublished data from the same
laboratory showed that administration of Aroclor 1221 at doses
<25 mgAg/day was not fetotoxic to rabbits (Villeneuve et al. 1971).
Doses of 0, 6.25, 12.5, 25, 50, or 100 mg/kg/day of Aroclor 1254
were administered by gavage on days 6 to 15 of gestation to rats
(Villeneuve et al. 1971). Average pup weights were reduced at
100 mg/kg/day, although total litter weight (average weight times number
of fetuses) did not differ from controls. There were no skeletal or
visceral abnormalities or effects on conception, resorptions, litter
size or number, or average litter weight in any of the treated groups.
In other rat studies with Aroclor 1254 (purity not reported), reduced
average fetal weight per litter (Spencer 1982) and reduced pup survival
and body weight at weaning (Linder et al. 1974) resulted from
100 mg/kg/day gavage exposure on days 6 or 7 to 15 of gestation.
Collins and Capen (1980a) fed diets containing 0, 50, or 500 ppm
Aroclor 1254 (purity not reported) to groups of 15 female Osborne-Mendel
rats throughout pregnancy and lactation. There was a statistically
significant (P < 0.001) reduced litter size in the 500-ppm groups
compared with controls. Statistically significant decreases in pup body
weight were observed at 50 and 500 ppm in 21-day-old pups, but not at
7 or 14 days or at parturition. Ultrastructural lesions in thyroid
follicular cells and reduction in serum levels of thyroid hormone
(thyroxine and triiodothyronine) occurred in the neonatal and weanling
rats at 50 and 500 ppm. Although pups are not usually examined for
effects on the thyroid in developmental studies, the observation of
thyroid effects in the neonates can be considered a fetotoxic effect
because the thyroid is a target organ of Aroclor 1254 toxicity. Assuming
that a rat consumes a daily amount of food equal to 5% of its body
weight (EPA 1986a), the 50- and 500-ppm levels are equivalent to doses
of 2.5 and 25 mg/kg/day, respectively; therefore, 2.5 mg/kg/day is the
LOAEL for fetotoxicity in rats.
Haake et al. (1987) reported that treatment of pregnant C57BL/6
mice with Aroclor 1254 by gavage at 244 mg/kg on day 9 of gestation did
not result in any fetuses with cleft palate.
Groups of eight female monkeys were maintained on diets containing
0, 0.25, or 1.0 ppm of Aroclor 1016 (free of PCDF) in the diet for
approximately 7 months prior to mating and during pregnancy (total
duration 87 ± 9 weeks) (Barsotti and Van Miller 1984). Mean birth weight
in the 1.0-ppm group was significantly (P < 0.01) less than controls,
but head circumference and crown-to-rump length were unaffected. All
females conceived, carried their infants to term, and delivered viable
offspring. More pronounced fetotoxic effects (early abortions or
resorption, stillbirths, and/or reduced birth weight), lengthened
menstrual cycles, and lowered serum progesterone levels occurred in
monkeys exposed to 2.5 or 5.0 ppm Aroclor 1248 (purity not reported) in
similarly designed studies (Allen and Barsotti 1976; Allen et al. 1979,
1980).
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70 Section 4
4.3.3.3 Dermal
Pertinent data were not located in the available literature.
4.3.3.4 General discussion
Comprehensive teratological examinations have not been conducted;
however, the above studies and others (EPA 1988a) indicate that Aroclors
were not teratogenic in rats and nonhuman primates when tested via the
oral route during the critical periods of organogenesis at doses that
produce fetotoxicity and/or maternal toxicity. Although fetotoxicity of
Aroclors is documented in several species of animals, the possibility
that contaminants (e.g., PCDFs) may be responsible for the effects
should be recognized.
The reports of reduced birth weight, gestational age, and
behavioral effects in infants of mothers with environmental and
occupational exposure to PCBs are inconclusive for the reasons indicated
in Sects. 4.3.3.1 and 4.3.3.2, but provide suggestive evidence for PCB-
related developmental effects in humans. Infants born to mothers who
were exposed to Kaneclor PCBs during the Yusho incident had signs of
toxicity and delayed development (e.g., abnormal skin pigmentation,
ocular discharge, small size), but no developmental abnormalities (EPA
1988a, Miller 1985). These effects did not persist. As discussed earlier
in this profile, the Yusho incident was a unique event in which effects
may not be related entirely to PCBs given that dibenzofurans were also
present.
Higher concentrations of PCBs in breast milk than in cord serum and
in suckling animals than in fetuses have led some investigators to
assume that postnatal lactation exposure poses a greater threat to
infants than intrauterine exposure. Jacobson et al. (1985) indicated
that this assumption may be inappropriate because fetuses may be
particularly sensitive to toxic insult due to factors such as lack of
protective barriers (i.e., blood-brain) and metabolizing capacities that
are found postnatally. That intrauterine exposure may be more harmful
than postnatal exposure is suggested by the results of the Jacobson et
al. (1985) study, which indicated that behavioral effects were
correlated more with prenatal exposure (cord serum PCBs) than with
exposure via breast milk.
4.3.4 Reproductive Toxicity
Data for reproductive effects in animals were available only for
oral exposure.
Groups of 12 female and 4 male mink were maintained on diets that
provided 0, 1, 5, or 15 ppm Aroclor 1254 (purity not reported) for
4 months and were mated (Aulerich and Ringer 1977). Dose-related
impaired reproduction (reduced number of females whelped and reduced
kit/female ratio) occurred at *5 ppm, with total inhibition of
reproduction at 15 ppm. These effects were also produced at 2 ppm
Aroclor 1254 in a similarly designed single-dose level study; however,
these effects did not appear to result from adverse effects on
spermatogenesis (Aulerich and Ringer 1977). Complete reproductive
failure occurred in mink exposed to >5 ppm Aroclor 1242, and
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Toxicological Data 71
Aroclor 1016 reduced but did not completely eliminate mink reproduction
at 20 ppm (Bleavins et al. 1980). The rat appears less sensitive, with
fetal mortality and maternal toxicity reported after daily consumption
for 9 weeks of Aroclor 1254 at a level of 6.4 mg/kg/day (Baker et al.
1977). The purity of the Aroclors was not reported.
Rats were exposed to 0, 1, 5, 20, 100, or 500 ppm of Aroclor 1254
(purity not reported) in the diet in one- and two-generation
reproduction studies (Linder et al. 1974). Reduced litter sizes occurred
in the Fib and F2 generations at >20 ppm.
In longer-term studies (Allen et al. 1979, 1980; Barsotti et al.
1976), monkeys were exposed to Aroclor 1248 in the diet at levels of
2.5 and 5.0 ppm for 18 months. Maternal toxicity that included
lengthened menstrual cycles was observed. At the high-dose level, there
was nearly complete inhibition of reproduction, while at the low-dose
there were early abortions and fetal resorptions, although some live
births did occur. Although this indicates that the monkey was very
sensitive to the reproductive toxicity of Aroclor 1248, it should be
noted that chemical analyses indicated that the PCBs were contaminated
with approximately 1.7 ppm of PCDFs, which may have contributed to the
observed toxicity.
Reproductive effects resulting from higher oral doses of Aroclor
prior to and during gestation include prolonged estrous cycle and
decreased sexual receptivity in rats (Brezner et al. 1984), reduced
conception rate in mice (Welsch 1985), and reduced litter size in rats
(Linder et al. 1974). Lactation exposure produced decreased reproductive
capacity in male rats (Sager 1983) and premature vaginal opening and
delayed first estrus in female rats (Brezner et al. 1984).
4.3.5 Genotoxicity
4.3.5.1 Human
No data were located in the available literature.
4.3.5.2 Animal
Results of mutagenicity assays with PCBs in in vitro systems are
summarized in Table 4.3. Results of studies using PCB mixtures other
than Aroclors are included to provide additional information. PCBs gave
generally negative results in Salmonella typhimurium, with and without
metabolic activation.
PCBs gave generally negative results in in vivo assays with rats
and mice (Table 4.4). Equivocal results (chromosomal aberrations) were
obtained in ring dove (Streptopchia risoria) embryos from doves fed
Aroclor 1254 at 10 ppm in the diet (Peakall et al. 1972).
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72 Section 4
Table 4.3. Genotoxicity of PCBs in vitro
End point
Species Result
(test system) with activation/without activation
References
Gene mutation
Chromosomal
aberrations
Salmonella
typhimurium
-I-
Chinese hamster
V79 cells
Human lymphocytes
Schoeny et al. 1979,
Schoeny 1982,
Heddle and Bruce 1977,
Wyndham et al. 1976,
Safe 1980,
Harbison 1986,
Bruce and Heddle 1979
Hattula 198S
Hoopingarner et al. 1972
Table 4.4. Genotoxicity of PCBs in vivo
End point
Species
(test system)
Result"
References
Chromosomal
aberration
Drosophila melanogaster
Ring dove
(Streptopchia risoria)
Chicken
Mouse
Rat
Micronucleus test Mouse
Sperm abnormality Mouse
Dominant lethal Mouse
Nilsson and Ramel 1974
Peakall et al. 1972
Blazak and Marcun 1975
Watanabe and Sugahara 1981
Green et al. 197Sa
Garthoff et al. 1977,
Dikshith et al. 1975
Bruce and Heddle 1979
Bruce and Heddle 1979
Green et al. 1975b,
Keplinger et al. 1971,
Calandra 1976
" negative
± equivocal
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Toxicological Data 73
4.3.6 Carclnogeniclty
4.3.6.1 Inhalation
Human. Two brief reports of a study of 31 research and development
employees and 41 refinery plant employees at a New Jersey petrochemical
facility (Bahn et al. 1976, 1977; Lawrence 1977) and an update of the
same study (NIOSH 1977b) are available. Aroclor 1254 had been used at
the plant during i 9-year period ending in the late 1950s. Malignant
melanomas were found in 2 of the 31 research and development workers and
1 of the 41 refinery plant workers; the incidence in the research and
development workers was significantly (P < 0.001) greater than expected.
NIOSH (1977b) found that there were 8 cancers in the total study
population (5.7 expected). Of these 8 cancers, 3 were melanomas and 2
were pancreatic cancer; these were significantly different from
calculated expectations (data not reported). The data from this study
should be regarded as inconclusive because PCB exposure was not
quantified, exposure to other potential and known carcinogens was not
evaluated although believed to be present, the number of cases and the
cohort size are small, and the expected cancer rates were based on U.S.
population data rather than on local rates.
Davidorf and Knupp (1979) found no relationship between possible
PCB exposure and increased annual occurrence of ocular melanoma in Ohio
during 1967-1977.
Brown and Jones (1981) conducted a retrospective cohort mortality
study of 2,567 workers who had completed at least 3 months of employment
during the years 1940-1976 (39,018 total person-years) in two capacitor
factories where PCBs were used. Aroclor 1254 was used first, but this
changed during the years to Aroclor 1242 and finally to Aroclor 1016.
Historical exposure data were not available, but personal TWA PCB
(Aroclor 1016) concentrations in 1977 ranged from 24-393 mg/m3 at plant
1 and 170-1,260 /ig/m3 at plant 2. Mortality from all causes and all
cancers was lower than expected. Excess mortality was noted for liver
cancer (3 observed deaths versus 1.07 expected) and rectal cancer (4
observed versus 1.19 expected), but neither excess was statistically
significant. There were no deaths due to malignant melanoma.
An unpublished update of the Brown and Jones (1981) study evaluated
an additional 7 years of follow-up (Brown, 1986) . Mortality from all
causes and all cancers was still lower than expected, but a
statistically significant (? < 0.05) excess risk of cancer of the liver
and biliary passages (5 observed versus 1.9 expected) was found. Four of
the 5 deaths due to liver cancer occurred in women who were employed in
plant 2; female employees at plant 2 contributed 41% of the total
person-years to the analysis. The author indicated that the liver cancer
can only be associated tentatively with PCB exposure because of the
small number of deaths and other limitations of the study.
Gustavsson et al. (1986) performed a cohort study of 142 male
Swedish capacitor-manufacturing workers who had been exposed to PCBs for
an average of 6.5 years between 1965 and 1978. Airborne PCB levels
measured in 1973 were 0.1 mg/m3. It is not clear if this level
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74 Section 4
represents an average for 1965-1978. Skin contamination had occurred in
some of the workers. Seven cancers had occurred in these workers, which
was in agreement with national statistics. One person had two rare
tumors, a slow-growing mesenchymal tumor and a malignant lymphoma. The
authors concluded that this study did not indicate any excess mortality
or cancer incidence among PCB workers, but that such effects could not
be ruled out because of the small cohort and relatively short follow-up
period.
Bertazzi et al. (1987) conducted a retrospective prospective
mortality study of 544 male and 1,556 female workers who were engaged in
the manufacture of PCB-impregnated capacitors in an Italian plant during
1946-1982. The workers were employed for a minimum of 1 week between
1946-1978 (41,010 person-years total) and examined for the period 1946-
1982. PCB mixtures containing 54% chlorine (Aroclor 1254 and Pyralene
1476) were used until 1964; these were progressively replaced by
mixtures containing 42% chlorine (Pyralene 3010 and 3011) until 1970,
when only Pyralene 3010 and 3011 were used. The maximum quantities of
PCBs were used in 1967-1968 and the use of PCBs has been abandoned
completely since 1980. Area samples taken in 1954 and 1977 showed air
PCB concentrations ranging from 5,200-6,800 0g/m3 (Aroclor 1254) and
48-275 mg/m3 (Pyralene 3010). Measurements of unspecified PCBs on
workers' hands in 1977 and 1982 showed concentrations ranging from 0.3-
9.2 pg/cmz and 0.09-1.5 jig/cm2, respectively. Mean blood concentrations
determined in 1977 and 1982 from the same 37 workers were 282.8 and
202.8 ppb for 54% chlorine PCBs. respectively, and 142.8 and 42.9 ppb
for 42% chlorine PCBs, respectively. Relatively few deaths were recorded
by 1982 [30 males (5.5%) and 34 females (2.2%)]. Overall mortality was
not significantly different from expected in males when compared with
national or local rates but was significantly (? < 0.05) higher than
expected in females when compared with local rates. Mortality from all
cancers was significantly higher than expected in males when compared
with both national and local rates (14 observed versus 1.7 national and
2.2 local), and in females when compared with local rates (12 observed
versus 5.3 expected). Deaths from gastrointestinal tract cancer were
significantly Increased in the males when compared with national and
local rates, and deaths from hematologic neoplasms were increased in
both sexes but only significantly in females when compared with local
rates. Clear site-specific risks of cancer cannot be identified because
of the small number of cases and limited follow-up.
Animal. No data were located In the available literature.
4.3.6.2 Oral
Huaan. Appropriate data were not located in the available
literature. Information regarding cancer in people exposed to PCBs
during the Yusho Incident is discussed in Sect. 4.3.6.4.
Animal. Kimbrough et al. (1975) fed groups of 200 female weanling
Sherman rats diets containing 0 or 100 ppm Aroclor 1260 (purity not
reported). Aroclor treatment was discontinued 6 weeks before the rats
were killed at 23 months of age. Mean final body weights and body weight
gain were significantly (P < 0.001) reduced in the treated group, but
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Toxicological Data 75
food consumption In the two groups was comparable. Actual PCB intake in
the treated rats was 11.6 mg/kg/day during the first week of exposure,
6.1 mg/kg/day at 3 months, and 4.3 mg/kg/day at 20 months. Almost all
treated rats (170/184) exhibited a few to multiple tan nodules on the
surface of the liver and more on sectioning. Only one control animal had
gross abnormalities of the liver. Hepatocellular carcinomas were found
in 1/173 (0.58%) controls and 26/184 (14%) treated rats. Neoplastic
nodules were found in the livers of 0/173 controls and 144/184 treated
rats. The total incidence of neoplastic liver lesions was 1/173 (<1%) in
controls and 170/184 (92%) in treated rats.
In a shorter preliminary study, Kimbrough et al. (1972) exposed
groups of 10 male and female Sherman rats to 0, 100, 500, or 1,000 ppm
Aroclor 1254 (purity not reported) or 1260 in the diet for <1 year. No
neoplastic nodules or hepatocellular carcinomas were found.
Norback and Veltman (1985) fed a group of Sprague-Dawley rats
(70 per sex) a diet containing Aroclor 1260 (purity not reported) at a
concentration of 100 ppm for 16 months, and 50 ppm for an additional
8 months, followed by a control diet for 5 months. A control group
consisted of 63 rats per sex. In the treated rats examined after
18 months, 95% of the 47 females and 15% of the 46 males had
hepatocellular neoplasms. This indicated a gender-related effect. Among
treated females, 43/47 had trabecular carcinomas and/or adenocarcinomas,
and another 2 females had neoplastic nodules only. Two of 46 treated
males had trabecular carcinomas, and another 5 had neoplastic nodules.
Incidences of hepatocellular neoplasms in control rats were 0/32 males
and 1/49 females, the one female having a single neoplastic nodule. The
progression of hepatocellular lesions was as follows: centrolobular cell
hypertrophy at 1 month, foci of cell alteration at 3 months and areas at
6 months, neoplastic nodules at 12 months, trabecular carcinoma at
15 months, and adenocarcinoma at 24 months. The authors noted that while
the tumors met morphologic criteria for malignancy, they were relatively
unaggressive as they did not metastasize to distant organs or invade
blood vessels. Mortality was not affected, probably because of the late
appearance and slow growth of the tumors. Both treated and control rats
developed cholangioma, cystic cholangioma, and adenofibrosis, but the
incidence was greater in the treated group.
EPA (1988a) used the Norback and Weltnan (1985) study as the basis
for a carcinogenic risk assessment of PCBs using combined incidences of
neoplastic nodules and hepatocellular carcinomas. Because this study
demonstrated the progression of hepatocellular lesions through
neoplastic nodules to carcinomas, it provides justification for using
the combined incidences for quantitative risk assessment.
NCI (1978) exposed groups of 24 Fischer 344 rats per sex per dose
to 0, 25, 50, or 100 ppm Aroclor 1254 in the diet for 104 to 105 weeks.
Mean body weights of mid- and high-dose males and low-dose females were
below those of controls from week 10 onward. There was a significant
dose-related reduction in survival among treated males. There was a
significant dose-related trend in combined incidences of lymphomas and
leukemias in males, but incidences in each dose group were not
significantly different from matched controls. NCI (1978) concluded that
these tumors could not clearly be related to administration of Aroclor
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76 Section 4
1254. Hepatocellular adenomas and carcinomas were found In treated
groups but not controls (males: mid-dose 1/24, high-dose 3/24; females-
mid-dose 1/24, high-dose 2/24). Nonneoplastic hyperplastic nodules also
occurred at a high incidence in treated animals but not controls. The
tumor incidences were not significant, but the hyperplastic nodules
appeared to be treatment related. Adenocarcinomas were found in the
stomach, jejunum, or cecum of two treated males and two treated females,
and a carcinoma was found in one treated male. Although their incidence'
was not statistically significant, the low historical incidences of
these lesions suggest that they might have been treatment related. NCI
(1978) concluded that the high incidence of hepatocellular proliferative
lesions in male and female rats was related to treatment, but that
Aroclor 1254 was not carcinogenic in this bioassay. There was no attempt
to identify or quantitate impurities.
Morgan et al. (1981) reexamined the NCI (1978) data with respect to
gastric adenocarcinomas. Stomachs from rats used in that study were
available for further sectioning and examination. Incidences of focal
stomach lesions, mostly metaplasia, were 6. 10, 17, and 35% in rats
receiving 0, 25, 50, and 100 ppm, respectively. Adenocarcinomas were
found in six treated rats. When compared with incidences of stomach
adenocarcinomas in historical controls (1/3,548), the incidence 6/144
was significant at P < 0.001. The authors commented that adenocarcinoma
and intestinal metaplasia appeared to be related and might have the same
initiating mechanism. They concluded that Aroclor 1254 led to induction
of intestinal metaplasia and probably to induction of adenocarcinoma in
the glandular stomachs of F344 rats.
Ward (1985) also reexamined data from the NCI (1978) bioassay. He
noted that hepatocellular adenomas, carcinomas, and eosinophilic and
vacuolated hepatocellular foci usually occurred only in treated rats. It
appeared that eosinophilic hepatocellular foci and tumors arose de novo
rather than from naturally occurring basophilic foci. He suggested that
Aroclor 1254 induced or initiated these unique lesions rather than
promoted the growth of naturally occurring lesions. Ward (1985) also
discussed the intestinal metaplasia and adenocarcinomas in treated rats.
He noted that the metaplastic lesions were similar to those seen in
monkeys, but differed in being focal and singular, while monkey lesions
were diffuse. The appearance of the few metaplastic lesions in the
stomachs of controls was different from those in treated rats, which
resembled precancerous lesions induced by gastric carcinogens. Ward
(1985) concluded that the effects of PCBs on the glandular stomach of
rats should be studied further.
Kimbrough and Linder (1974) fed groups of 50 male Balb/cJ mice
diets containing 0 or 300 ppm Aroclor 1254 (purity not reported) for
11 months or for 6 months followed by a 5-month recovery period. Treated
mice had enlarged livers and adenofibrosis, a possible premalignant
lesion (EPA 1988a). Incidences of hepatomas were: 0/34 and 0/24 in two
control groups, 9/22 in the 11-month exposure group, and 1/24 in the
5-month exposure group. This study provided evidence of the potential
hepatocarcinogenicity of PCBs in mice.
Ito et al. (1974) observed hepatocellular carcinomas (5/12 mice)
and liver nodules (7/12) in dd mice fed 500 ppm of Kanechlors 500 for
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lexicological Daca 77
32 weeks. This study provides supporting evidence for the hepato-
carcinogenicity of PCB mixtures.
Because PCB mixtures are often contaminated with PCDFs, it is
possible that the carcinogenic response of some PCB mixtures is due to
or augmented by these contaminants. Schaeffer et al. (1984) fed male
Wistar rats diets containing 100 ppm Clophen A-30 (30% chlorines by
weight) or Clophen A-60 (60% chlorines by weight) until they died
(approximately 800 days). These PCB mixtures were reported to be free of
furans. The treated rats displayed better survival than did controls.
Hepatocellular carcinomas developed in 61% of the rats fed Clophen A-60.
Only 3% of the Clophen A-30 treated rats developed hepatocellular
carcinomas, while 89% had preneoplastic lesions. None of controls
developed hepatocellular carcinomas. This study demonstrates that PCB
mixtures free from contamination with furans elicit a carcinogenic
response.
4.3.6.3 Dermal
Human. Human exposures to PCBs via both the dermal and inhalation
routes are discussed under the inhalation data.
Animal. DiGiovanni et al. (1977) reported that Aroclor 1254
(purity not reported) showed weak initiator activity when applied to the
skin of CD-I mice as a single 0.1-mg dose, followed by promotion with
the phorbol ester TPA (12-0-tetradecanoylphorbol-13-acetate).
Interpretation of this study is confounded by the lack of a control
group treated only with TPA; TPA and other phorbol esters have been
shown to produce low incidences of skin tumors (Berry et al. 1978, 1979;
Van Duuren 1981). Berry et al. (1978, 1979) reported that Aroclor 1254
was not a skin tumor promoter in female CD-I mice that had been
initiated with dimethyIbenzanthracene (DMBA), nor did it produce tumors
when tested without DMBA initiation at a level of 0.1 mg administered
twice weekly.
4.3.6.4 General discussion
The study by Kimbrough et al. (1975) demonstrated the
hepatocarcinogenicity of Aroclor 1260 In female Sherman rats. A
preliminary experiment using smaller groups of animals of the same sex
and strain exposed for <1 year did not result in neoplastic nodules or
hepatocellular carcinomas (Kimbrough et al. 1972). These results suggest
that hepatocellular carcinomas caused by PCBs can be detected only in
long-term experiments at doses low enough to prevent interfering
toxicity (EPA 1985a). In addition, because the large long-term
experiment only produced a 14% incidence of carcinomas, relatively large
numbers of animals must be used to detect a significant increase in
tumor incidence. Similarly, the NCI (1978) rat study with group sizes of
24 rats per sex was considered not sensitive enough to identify as
significant an increase in tumor incidence of this magnitude (14%). The
NCI (1978) study found hepatocellular carcinomas in 2/24 (8%) male rats
fed 100 ppm Aroclor 1254. If incidences are expressed as the number of
animals with tumor per number of animals at risk, as is more commonly
done, the incidence is 2/20 or 10%. The 8 to 10% incidence is not
detected as statistically significant with group sizes of 24 rats, nor
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78 Section 4
would a 14% incidence, as was observed in Che Kimbrough et al (1975)
study, be detected as statistically significant. The studies of Aroclor
1260 and Clophen A-60 indicate that liver tumors induced by these 60%
chlorine PCB mixtures are relatively unaggressive, nonmetastasizlng and
not life-shortening, and that incidences of extrahepatic tumors are
decreased (Kimbrough et al. 1975, Schaeffer et al. 1984, Young 1985,
Norback and Ueltman 1985). In the studies by Kimbrough et al. (1975)
the rats were killed at 23 months of age. a substantial portion of their
life span, and apparently, there was no significant difference in
mortality between the control and treated group, although the treated
rats had very significant increased incidences of liver tumors compared
with controls. In the study by Norbeck and Veltman (1985), the treated
and control rats were maintained for a total of 29 months. While the
treated rats developed highly significant increased incidences of liver
tumors compared with controls, there was no effect on mortality. In the
study by Schaeffer et al. (1984), the rats were followed until they
died; the treated animals, which had increased incidences of liver
tumors, actually survived longer than the controls.
EPA (1985a) discussed the difficulties in using data from assays
with commercial PCB mixtures for quantitative risk assessment. The
composition of these mixtures is highly variable. Different lots of the
same Aroclor, while having the same average chlorine content, can differ
substantially in content of individual isomers. The metabolic and
pharmacokinetic behavior of the pure isomers varies greatly with the
degree and position of chlorine substituents. Analysis of an Aroclor
1254 lot indicated a predominance of pentachloro biphenyl isomers, whic*
are relatively rapidly metabolized and excreted. An Aroclor 1260 lot wa.
primarily hexa- and heptachloro isomers, which would be retained in
adipose and skin storage depots for long periods. These storage depots
might be considered effective removal of carcinogens from the target
organs or, conversely, a carcinogen pool capable of mobilization and
adding to target organ exposure. Different Aroclors administered at the
same dosage could result in completely different tissue-specific
exposure levels for the various pure Isomers and metabolites. A potency
estimate based only on administered dosage is therefore inappropriate.
EPA (1985a) concluded that the potency of any commercial PCB mixture is
probably higher than any estimate that would be derived by using dietary
levels of exposure as a basis for calculation.
EPA (1988b) concluded that the level of carcinogenic evidence in
rats and mice for some commercial PCBs (Aroclor 1260, Kanechlor 500, and
Aroclor 1254) constitutes a "sufficient" level of carcinogenic evidence
for PCBs in animals. The multiple studies with Aroclor 1260 and one
study with Clophen A-60 provide sufficient animal cancer evidence, and
the studies with Aroclor 1254. Kanechlor 500, and Clophen A-30 provide
limited animal cancer evidence. Taken collectively, this evidence, along
with an argument for a hypothesis that structure-activity relationship
provides a basis for recommending that PCB mixtures of any composition
should be regarded as having the potential to be probable human
carcinogens, is used to classify PCBs in the EPA veight-of-evidence
category Group B2 (EPA 1988b). The EPA (1988b) decision to regard all
PCBs as Group B2 compounds has uncertainty since it cannot be verified
with present knowledge.
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Toxicologies! Data 79
EPA (198Sa) selected the Kimbrough et al. (1975) study as the basis
for the earcinogeniclty risk assessment for PCBa. More recently, the
Norback and Veltman (1985) study was used for quantitative risk
assessment in EPA (1988a,b), which supersedes the aforementioned
assessment. The Norback and Weltman (1985) study was preferred because
the strain of rats used (Sprague-Dawley) has a low incidence of
spontaneous liver neoplasia, the duration of the study was for the life
span of the rats, and there was a sequential progression of liver
lesions to hepatocellular carcinomas.
Available epidemiological data do not indicate a consistent
tumorigenic effect among people exposed to PCBs. As indicated in
Sect. 4.3.6.1, occupational studies (Brown 1986, Bertazzi et al. 1987)
suggest possible carcinogenicity of PCBs by the inhalation route. A
statistically significant excess risk of liver cancer has been reported
in Yusho patients who were studied for a follow-up period of >16 years
(Amano et al. 1984, Kuratsune 1986). Because the excess cancer was found
in only one prefecture and the victims also consumed PCDFs and
polychlorinated quaterphenyls, these findings are considered to be
suggestive of a possible carcinogenic effect of PCBs by the oral route.
Because of the tentative nature of the inhalation and oral data, EPA
(1988b) has concluded that the evidence for carcinogenicity in humans is
inadequate but suggestive.
4.4 INTERACTIONS WITH OTHER CHEMICALS
Many of the interactive effects of PCBs with other chemicals are
related to the capacity of PCBs for enzyme induction. Therefore, the
effects of PCBs on toxicity of other compounds depend on the role of
oxidative metabolism in the toxicity of those compounds. Reported
effects of PCB pretreatment include Increased metabolism and excretion
of pentobarbital and decreased pentobarbital sleeping times (Chu et al.
1977, Villeneuve et al. 1972), increased mutagenicity of B(a)P (Button
et al. 1979), and increased hepatotoxlcity of halothane and vinylidene
fluoride (Sipes et al. 1978, Conolly et al. 1979).
Increased dietary ascorbic acid may protect against some of the
toxic effects of PCBs, such as altered enzyme activity and liver
histopathology, perhaps by Inhibiting lipid peroxidation (Chakraborty et
al. 1978, Kato et al. 1981). The exact mechanism is not known.
PCBs have had mixed effects on tumor development. Aroclor 1254
pretreatment protected mice from lung tumors but increased the number of
mice with liver tumors 18 months after administration of
N-nitrosodlmethylamine (Anderson et al. 1983). Pretreatment with Aroclor
1254 gave slight protection against skin tumor development in mice
initiated with 7,12-dimethylbenz(a)anthracene and promoted with TPA
(Berry et al. 1979). Makiura et al. (1974) reported that Kanechlor 500
inhibited hepatocarclnogenicity of 3*-methyl-4-dimethylaminoazobenzene,
N-2-fluorenylacetamide, and N-nitrosodiethylamine when administered
orally to rats. Nagasaki et al. (1975) found that Kanechlor 400 and 500
enhanced the hepatocarcinogeniclty of a-BHC In mice. PCBs promoted the
development of enzyme-altered foci or hyperplastlc nodules following
treatment with nltrosamines (Oesterle and Demi 1983, Pereira et al.
1982) or N-2-fluorenylacetamide (Tatematsu et al. 1979).
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80 Section 4
Birnbaum et al. (1985) found that 2,3,3',4,4-,5-hexachlorobiphenyl
but not 2,2',4,4',5,5'-hexachlorobiphenyl, when coadministered with
2,3,7,8-TCDD to mice during gestation resulted in a dose-related
enhancement of the TCDD-induced hydronephrosis in mouse fetuses, but
2,3,3',4,4',5-hexachlorobiphenyl alone caused hydronephrosis in'the
mouse fetuses. 2,2',4,4',5,5'-Hexachlorobiphenyl did not induce
hydronephrosis.
Haake et al. (1987) found that Aroclor 1254 antagonized the
teratogenicity of 2,3,7,8-TCDD in mice. In this study, treatment of
pregnant mice by gavage with Aroclor 1254 at 244 mg/kg on day 9 of
gestation followed by 2,3,7,8-TCDD at 20 mgAg on day 10 resulted in an
8.2% incidence of cleft palate. Treatment with 2,3,7,8-TCDD alone
resulted in a 62% incidence of cleft palate. Aroclor 1254 alone was not
teratogenic.
Bannister et al. (1987) found that Aroclor 1254 partially
antagonized the 2,3,7,8-TCDD-induced microsomal enzyme induction and
immunotoxicity in mice.
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81
5. MANUFACTURE. IMPORT. USE, AND DISPOSAL
5.1 OVERVIEW
PCBs are no longer produced or used in the United States; however,
many of the transformers and capacitors which were produced with PCBs,
and contain PCBs, are still in service. Therefore, these products
constitute a potential source of exposure to the environment and to
humans. Disposal of PCB materials is controlled by federal regulations.
5.2 PRODUCTION
PCBs have been commercially produced in the United States since
1929. Annual U.S. production of PCBs peaked in 1970 when 85 million
pounds were produced. It was estimated that approximately 1,000 million
pounds of PCBs had been sold in North America by 1970. Manufacture of
PCBs (Aroclors) in the United States was terminated in October 1977
because these products accumulated and persisted in the environment and
because of their toxic effects. Monsanto, the sole U.S. manufacturer at
that time, had been producing Aroclors 1016, 1221, 1242, and 1254. In
1974, Monsanto produced Just over 40 million pounds of the Aroclor
mixtures. Production had been approximately 40 million pounds annually
since 1971. Monsanto produced PCB Aroclor products at a facility in
Sauget, Illinois, but production was stopped in October 1977. Of the
total PCBs sold in the United States since 1970, over 98% were Aroclor
1260, 1254. 1248, 1242. 1232, 1221, and 1016 and less than 2% were
Aroclor 1268 and Aroclor 1262. Therefore, 98% of PCBs sold in the United
States since 1970 have been covered in this document (IARC 1978, Hatton
1979, Durfee 1976, EPA 1976).
The Aroclors were prepared industrially by the chlorination of
biphenyl with anhydrous chlorine in the presence of a catalyst such as
iron filings or ferric chloride. The degree of chlorination, which
determined which Aroclor was produced, was controlled by the anhydrous
chlorine contact time in the reactor (EPA 1976).
5.3 IMPORT
Imports of PCBs through principal U.S. custom districts in recent
years have been reported as follows (USITC 1978, 1979, 1980, 1982):
Import volume
Year (Ib)
1981 11.000
1979 357.147
1978 483.074
1977 280.867
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82 Section 5
No data were located to indicate that PCBs have been imported after
1981.
Section 6(e)(3)(A) of TSCA (Pub. L. 94-469, 90 stat. 2003,
15U.S.C.2601 et seq) prohibits all manufacture and importation of PCBs
as of January 1, 1979. As of January 2, 1979, EPA announced that
companies that had filed petitions for exemptions from the PCB
manufacturing/importation ban could continue the manufacturing or
importation activity until EPA has acted on the application petition
(EPA 1979).
5.4 USES
A thorough review of PCB use in the United States can be found in
EPA (1976). By 1974, all U.S. use of PCBs was in closed systems for the
production of capacitors and transformers. As of 1976, 70% of Monsanto's
domestic sales of Aroclors was used in capacitor production and 30% in
transformer production. Aroclors are no longer used in the production of
capacitors and transformers; however, many of the devices manufactured
with Aroclors are still in service today. The life expectancy of
transformers containing PCBs is >30 years, and the life expectancy of
capacitors can range from 10 to >20 years, depending upon electrical
application. PCBs were used in capacitors and transformers because of
their excellent dielectric properties and fire resistance. Production of
a large capacitor involved filling the capacitor with the Aroclor oil
(typically over 2 to 3 Ib of PCB) through a small hole and then sealing.
Transformers were similarly filled, but may contain many times the
amount of PCBs, depending on size. As of 1976, only 5% of the
transformers produced in the United States were filled with PCBs, but
95% of the capacitors used PCBs (Durfee 1976). As of 1981, an estimated
131,200 PCB transformers were still in service in the United States,
representing approximately 1% of all operational transformers (Orris et
al. 1986). PCBs (Aroclors 1260 and 1262) have been used as a slide-
mounting medium for microscopes (IARC 1978) and are still used
occasionally for this purpose since this use has been exempted from
federal use restrictions.
5.5 DISPOSAL
On April 18. 1978, regulations became effective in the United
States concerning the storage and disposal of PCBs. These regulations
specified incineration of the waste or contaminated material as the only
acceptable method of PCB disposal unless, if this method is not
possible, clearance is obtained from the EPA to dispose of the materials
in another way. In March 1983, the EPA issued a procedural amendment to
the PCB rule to enable new disposal technologies to receive approval on
a nationwide basis. At present, EPA's PCB disposal rules typically
require that various types of PCBs and PCB materials be disposed of in
chemical-waste landfills or destroyed in high-temperature Incinerators
or high-efficiency boilers. The disposal rules are published in the July
1984 Code of Federal Regulations, 40CFR, Part 761 (Kokoszka and Flood
1985, Hatton 1979).
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83
6. ENVIRONMENTAL FATE
6.1 OVERVIEW
At present, the major source of PCB exposure in the general
environment appears to be environmental cycling of PCBs previously
introduced into the environment. This cycling process Involves
volatilization from ground surfaces into the atmosphere with subsequent
removal from the atmosphere via wet/dry deposition and then
revolatilization. The environmental persistence of PCBs generally
increases with an increase in the degree of chlorination of the
congener. The Aroclors with a high degree of chlorination (1248, 1254,
and 1260) are resistant to biodegradation and appear to be degraded very
slowly in the environment. The chemical composition of the original
commercial Aroclor mixtures which were released to the environment has
changed over time since the individual congeners degrade and partition
at different rates. Reviews of the environmental fate processes of PCBs
are available (EPA 1988a, Leifer et ml. 1983, Callahan et ml. 1979).
6.2 RELEASES TO THE ENVIRONMENT
Since the Aroclors are no longer produced or used in the production
of new products in the United States, industrial effluent discharges
from production sources no longer occur. Current sources of PCB release
to the environment include releases from landfills containing
transformers, capacitors, and other PCB wastes; waste incineration of
PCB materials; spills; and improper (or illegal) disposal to open areas
(Weant and McCormick 1984, Murphy et al. 1985). In addition, explosions
or overheating of transformers containing PCBs may release significant
amounts of these materials into the local environment.
PCB emissions from landfills and incinerator stacks have been
monitored (Murphy et al. 1985). Landfills are expected to be a
continuous source of PCB release into the atmosphere because methane and
carbon dioxide, which are generated from anaerobic degradation of
organic waste, are released and expected to carry PCBs and all other
volatile compounds with them. Incinerator stacks are expected to be a
source of PCBs, which would volatilize in the upper levels of the
incinerator before combustion occurred, because PCBs are resistant to
oxidation but reasonably volatile. This monitoring has indicated that
the amount of PCBs released from these sources (10-100 kg/year from
landfills and 0.25 kg/stack/year for incinerators) may not be
significant when compared to the quantities of PCBs estimated to cycle
annually through the atmosphere over the U.S. (900,000 kg/year).
Atmospheric fallout and washout have been Identified as nonpolnt
sources of PCB exposure to the environment (Kleinert 1976, Weant and
McCormick 1984, Swackhamer and Armstrong 1986, Larsaon 1985). Although
-------
84 Section 6
additional research is required for a definitive answer, evidence
suggests that the current major source of PCB release to the environment
is an environmental cycling process (Swackhamer and Armstrong 1986,
Larsson 1985, Murphy et al. 1985). This cycling process involves
volatilization of PCBs from bodies of water or from soil surfaces into
the atmosphere. Once in the atmosphere, the PCBs are returned to earth
via washout/fallout where the cycle is subsequently repeated with
revolatilization. Since the volatilization and degradation rates of PCBs
vary among the congeners present, this cycling process causes an
alteration of the PCB ratio in water and air relative to the original
source.
6.3 ENVIRONMENTAL FATE
6.3.1 Transport and Partitioning
In water, adsorption to sediments or other organic matter is a
major fate process for the PCBs (EPA 1988a, Callahan et al. 1979).
Experimental and monitoring data have shown that PCB concentrations are
higher in sediment and suspended matter than in the associated water
column. Based on their water solubilities and octanol-water partition
coefficients, the lower chlorinated components of the Aroclors will sorb
less strongly than the higher chlorinated isomers. Although adsorption
can immobilize PCBs for relatively long periods of time in the aquatic
environment, resolution into the water column has been shown to occur on
an environmental level (Swackhamer and Armstrong 1986. Baker et al.
1985). The substantial quantities of PCBs contained in aquatic sediments
can therefore act as an environmental sink for environmental
redistribution of PCBs. Environmental redistribution from aquatic
sediments should be most important for the PCBs contained in the top
layers of the sedimentary deposit. PCBs reaching the lower layers of
sedimentary deposits may be effectively sequestered from environmental
redistribution.
Volatilization is also an important environmental fate process for
the PCBs that exist in natural water in the dissolved state. The values
of the estimated Henry's law constants for the Aroclors (although they
occur as a mixture in natural water) (see Table 3.2) are indicative of
significant volatilization from environmental waters (Lyman et al.
1982). A study conducted on Lake Michigan has Indicated that
volatilization may be the major removal mechanism of PCBs from lakes
(Swackhamer and Armstrong 1986). Strong PCB adsorption to sediment,
however, significantly decreases the rate of volatilization, with the
higher chlorinated Aroclors having longer volatilization half-lives than
the lower chlorinated Aroclors (EPA 198Sa). However, eventual resolution
of PCBs from sediment into the water column can then result in
volatilization.
The low water solubility, high octanol-water partition coefficients
(see Chapter 3) of the PCBs and demonstrated strong adsorption of PCBs
to soils and sediment (EPA 1988a, Callahan et al. 1979, Sklarew and
Girvtn 1987) indicate that significant leaching should not occur in soil
under most conditions. The tendency of the lover chlorinated PCBs to
leach will be greater than the highly chlorinated PCBs. In the presence
-------
Environmental Face 85
of organic solvents, PCBs can leach significantly in soil (Griffin and
Chou 1981).
Organics having vapor pressures >10'4 mm Hg should exist almost
entirely in the vapor phase in the atmosphere, while organics having
vapor pressures <10"8 mm Hg should exist almost entirely in the
particulate phase (Eisenreich et al. 1981). The vapor pressures of the
Aroclors (see Table 3.2) indicate that they should therefore exist
primarily in the vapor phase in the atmosphere. Monitoring data have
shown that between 87 and 100% of the PCBs in air are operationally in
the vapor phase (Eisenreich et al. 1981). The tendency of PCBs to adsorb
to particulates will increase as the degree of chlorination increases.
PCBs in the atmosphere are physically removed by wet and dry
deposition (Eisenreich et al. 1981). Dry deposition occurs only for the
PCBs associated in the particulate phase. The PCB concentration of rain
anywhere in the world may typically range between 1 and 250 ng/L
(Eisenreich et al. 1981), which is an indication of the importance of
wet deposition.
6.3.2 Transformation and Degradation
The ability of PCBs to be degraded or transformed in the
environment is dependent upon the degree of chlorination of the biphenyl
molecule (EPA 1988a, Leifer et al. 1983, Callahan et al. 1979). In
general, the persistence of PCB congeners increases as the degree of
chlorination increases.
In the atmosphere, the vapor phase reaction of PCBs with hydroxyl
radicals (which are photochemically formed by sunlight) may be the
dominant transformation process. The estimated half-lives for this
reaction in a typical atmosphere with various PCB isomers are as follows
(EPA 1987b): monochlorobiphenyl, 12.9 days; dichlorobiphenyl, 27.8 days;
trichlorobiphenyl, 1.43 months; tetrachlorobiphenyl, 3.1 months;
pentachlorobiphenyl, 4.75 months; hexachlorobiphenyl, 10.3 months; and
heptachlorobiphenyl, 1.31 years.
In the aquatic environment, transformation processes such as
hydrolysis and oxidation do not significantly degrade PCBs (Mabey et al.
1981; Callahan et al. 1979). Photolysis appears to be the only viable
chemical degradation process in water; however, sufficient experimental
data are not available to determine its relative rate or importance in
the environment (Leifer et al. 1983).
Reviews of the biodegradability of PCBs are available (EPA 1988a,
Leifer et al. 1983). Biodegradation rates depend on a number of factors,
such as the amount of chlorination, concentration, type of microbial
population, available nutrients, and temperature; therefore, the rates
are highly variable. However, the results generally show that mono-,
di-, and trichlorinated biphenyls (major components in Aroclors 1221 and
1232) biodegrade relatively rapidly; tetrachlorinated biphenyls (major
components in Aroclors 1016 and 1242) biodegrade slowly; and the higher
chlorinated biphenyls (major components in 1248, 12S4, and 1260) are
resistant to biodegradation. In addition to the degree of chlorination,
chlorine positions on the biphenyl ring appear to be important in
determining the biodegradation rate. For example, PCBs containing all of
-------
86 Section 6
the chlorines on one ring are degraded faster than PCBs containing the
chlorines distributed between both rings, and PCBs containing chlorines
in the ortho positions are more resistant (Leifer et al. 1983). A study
of subsurface aquatic sediments has shown that PCBs containing chlorines
in the para positions are preferentially biodegraded as compared to
other ring positions (Brown et al. 1987). This study of subsurface
sediments, primarily from spill sites, has also shown that the higher
chlorinated congeners are biotransformed by a reductive dechlorination
to lower chlorinated PCBs which are biodegradable by aerobic processes.
This is important since PCBs in soil systems or in aquatic sediments
have not been shown to degrade by processes other than biodegradation.
Therefore, biodegradation is probably the ultimate degradation process
in soils and in sediments.
A summary of experimentally determined bioconcentration factors of
various Aroclors (1016, 1248, 1254, and 1260) in aquatic species (fish,
shrimp, oyster) has found Aroclor bioconcentration factors ranging from
26.000 to 660,000 (Leifer et al. 1983).
-------
87
7. POTENTIAL FOR HUMAN EXPOSURE
7.1 OVERVIEW
PCBs partition significantly from water to aquatic organisms such
as fish and can result in extremely high bioconcentration factors.
Consumption of contaminated fish then results in human exposure to PCBs.
Consumption of fish has been identified as a primary route of human
exposure to PCBs. The general population is also exposed, on a continual
basis, to PCB levels in the breathable air. PCBs have been found in at
least 216 of 1,177 sites on the National Priorities List (View 1989). A
review of environmental PCB monitoring data is available (EPA 1988a).
7.2 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT
7.2.1 Air
Eisenreich et al. (1981) completed the following list of typical
atmospheric concentrations of PCBs:
Concentration range
Location (ng/m3) Mean
Urban
Rural
Great Lakes
Marine
Remote
0.5 to 30
0.1 to 2
0.4 to 3
0.05 to 2
0.02 to 0.5
5-10
0.8
1
0.5
0.1
These values were derived from a large volume of monitoring data
reported in the literature.
Ambient atmospheric PCB concentrations of 7.1 and 4.4 ng/m3 were
detected in Boston. Massachusetts, and Columbia, South Carolina,
respectively, during the summer of 1978 (Bidleman 1981). These
concentrations are a composite for Aroclors 1016, 1242, and 1254.
Analysis of ambient air in Antarctica between 1981 and 1982 found PCB
levels of 0.02 to 0.18 ng/m3 (Tanabe et al. 1983).
The average PCB concentration (Aroclors 1242 and 1260) emitted from
gas vents at a hazardous waste landfill in North Carolina was found to
be 0.126 mg/m3 (Lewis et al. 1985). PCB concentrations of 0.01 to
1.5 ppm were detected in the fly ash from five municipal incinerators
operating under different technological and working conditions (Morselli
et al. 1985). Stack effluents from several midwest municipal refuse and
sewage incinerators contained PCB levels of 300 to 3,000 ng/m3 (Murphy
et al. 1985). The total PCB concentration measured in the flue gas
effluent from a municpal waste incinerator in Ohio was 260 ng/m'
(Tiernan et al. 1983). PCBs were detected in effluents from combustion
-------
88 Section 7
of coal and refuse at Ames, Iowa, at levels of 2 to 10 ng/m3 (EPA
1988a).
The average adult male inhales approximately 20 m3 of air per day.
Assuming the breathable outdoor air at a typical urban location contains
an average PCB concentration of 5 ng/m3, the average daily intake via
inhalation would be 100 ng. This estimate pertains to background levels
of PCBs in outdoor air. As reported in Sect. 7.4 (populations at high
risk), PCB levels in certain indoor air may be an order of magnitude
higher than in outdoor air.
7.2.2 Water
The concentration of PCBs in the open waters of the oceans can be
an indication of the environmental background level in water.
Concentrations reported for various seawaters include 0.04 to 0.59 ng/L
in the north Pacific, 0.035 to 0.069 ng/L in the Antarctic, and 0.02 to
0.20 ng/L in the north Atlantic (Tanabe et al. 1983, 1984; Giam et al.
1978). PCB concentrations of 0.3 to 3 ng/L, which are higher than the
seawater levels reported above, have been detected in seawater from the
North Sea; however, the seawaters sampled were receiving an
anthropogenic influence (Boon and Duinker 1986).
Mean PCB concentrations of 0.63 to 3.3 ng/L were detected in the
waters of western Lake Superior during 1978 to 1983 monitoring (Baker et
al. 1985). Mean levels of 3.0 to 9.0 ng/L (1974 to 1976) and 0.49 to
17.15 ng/L (1979 to 1981) were found in the water columns of Lake
Michigan and Lake Huron, respectively (Rodgers and Swain 1983). Analysis
of water from eight sites in Galveston Bay resulted in an average PCB
level of 3.1 ng/L between 1978 and 1979 (Murray et al. 1981). Thirty-two
of 163 wells monitored in industrialized areas of New Jersey were found
to contain PCB levels ranging from 60 to 1,270 ng/L (EPA 1988a). Mean
PCB levels of 25 to 38 ng/L were detected in waters collected from 11
agricultural watersheds in Ontario during 1975 to 1977 (Frank et al.
1982). A discussion of a number of PCB monitoring studies conducted on
the Hudson River can be found in EPA (1988a).
Although PCBs are widespread in the aquatic environment, their low
solubility generally prevents them from reaching high concentrations in
drinking water supplies (EPA 1980a). The National Organic Monitoring
Survey (NOMS) was conducted by the EPA to determine the frequency of
occurrence of specific organic chemicals (including PCBs) in finished
water supplies of 113 cities nationwide (EPA 1988a). Data from the three
phases (referred to as NOMS I, II, and III) of the study were collected
between March 1975 and January 1977. PCBs were not found in groundwater
supplies sampled in NOMS I (minimum quantifiable limit of 0.12 ppb).
Only a single finished groundwater sample in each of NOMS I and II
contained detectable levels of PCBs; the concentration of each was
reported to be 0.1 ppb (detection limits of 0.1 to 0.2 ppb). PCBs were
detected in two finished surface water supplies in each of NOMS I and II
and in one surface water in NOMS III; the concentrations of the five
positive samples ranged from 0.1 to 1.4 ppb. A total mean PCB level of
0.12 to 0.8 ppb was found in tap water from the Vaterford Water Co.
(Hudson River source) In 1976 and 1977 (EPA 1988a, Kim and Stone n.d.).
-------
Potential for Human Exposure 89
7.2.3 Soil
An analysis of 99 soil samples from rural and urban sites
throughout Great Britain was conducted to determine background levels of
PCBs in British soils (Greaser and Fernandes 1986). PCBs were identified
in all samples within the range of 2.3 to 444 ppb (MgAg)- The mean and
median values found for all samples were 22.8 and 7.2 ppb, respectively.
PCB levels ranging from 4.5 to 47.7 MgAg have been detected in soil
samples collected in the vicinity of incineration facilities in South
Wales and Scotland during 1984 to 1985 (Eduljee et al. 1985, 1986). An
analysis of Japanese soils detected PCB levels as high as 100 MgAg;
however, 40% of the samples had levels <10 MgAg (Greaser and Fernandes
1986).
PCB concentrations ranging from <1 to 33 ppb have been detected in
the soils of the Everglades National Forest in Florida (Requejo et al.
1979), which is consistent with the monitoring data from Great Britain.
Carey et al. (1979a) analyzed soils from 37 states in 1972 as part of
the National Soils Monitoring Program and found PCB in only 2 of 1,483
soil samples; however, the analytical technique used had a minimum
detectable limit of only 0.05 to 0.1 ppm, which was not low enough to
detect the mean and median levels reported in Great Britain. Carey et
al. (1979b) used the same analytical technique to analyze soils from
five U.S. urban areas (43-156 samples per site) in 1971; positive
detections were reported for three areas with PCB levels ranging from
0.02 to 11.94 ppm. The highest level (11.94 ppm) was detected in 1 of
55 samples from Gadsden Alabama.
PCB levels of 0.098 to 0.54 tag/kg have been detected in the
sediments from four remote high-altitude lakes in the Rocky Mountain
National Park (Heit et al. 1984), which indicates levels of PCBs that
can accumulate in sediments from natural deposition. Sediment core
samples from the Milwaukee harbor, which has received industrial
effluents of PCBs, have been found to contain levels of 1.03 to 13.4
mgAg (Christensen and Lo 1986). Analysis of sediments from 13 selected
streams in the Potomac River Basin found a maximum PCB level of
1.2 mgAg in one scream (Feltz 1980). In seven of the streams, zero or
trace amounts of PCBs were detected, but the rest contained 10-80 MgAg-
Upper sediment layers from the Hudson River and New York Harbor in 1977
contained Aroclor 1254 levels of 0.56 to 1.95 ppm and Aroclor 1242
levels of 3.95 to 33.3 ppm (Bopp et al. 1982). Analysis of surficial
sediments from the Great Lakes and various associated waters found
Aroclor 1254 levels of 2.5 to 251.7 ng/g, with the higher levels
detected in Lake Erie (Thomas and Frank 1981). An average Aroclor 1260
concentration of 120 ng/g has been detected in sediment samples from
eight sites along the coast of Maine (Ray et al. 1983).
7.2.4 Other
7.2.4.1 Foodstuffs
Table 7.1 lists the amounts of PCBs detected in raw domestic
agricultural commodities during fiscal years 1970 to 1976. These
commodities were analyzed as part of federal monitoring programs
conducted by the U.S. Food and Drug Administration (FDA) and the
-------
90 Section 7
Table 7.1. Aroclor residues in raw domestic agricultural
commodities for fiscal yean 1970-1976
Number of samples
Commodity analyzed
Fish
Shellfish
Eggs
Red meat6
Poultry
Ruid milk
Cheese
2,901
291
2.303
15,200
11.340
4.638
784
Percent with
positive detections
46.0
18.2
9.6
0.4
0.6
4.1
0.9
Average
concentration
(ppm)"
0.892
O.OS6
0.072
0.008
0.006
0.067
0.011
"Average fall samples, both positive and negative.
'Fiscal years 1973-1976.
Source: Duggan et al. 1983.
-------
Potential for Human Exposure 91
U.S. Department of Agriculture. It appears from Table 7.1 that fish are
the primary foodstuff containing environmental background levels of
PCBs; additional fish monitoring data are cited below. The contamination
of fish is a consequence of the contamination of the aquatic environment
and resulting bioconcentration (EPA 1980a).
Since the early 1960s, the FDA has conducted the Total Diet
Studies, which have also been known as the Market Basket surveys. These
studies, conducted on an annual basis, analyze ready-to-eat foods
collected in markets from a number of cities nationwide to determine the
intake of selected contaminants in the American diet. Table 7.2 presents
the recent results of the Total Diet Studies with respect to PCBs. Since
the mid-1970s, individual diets for adult males (19 years old), infants,
and toddlers have been analyzed. Assuming that the average adult male
weighs 70 kg and that the estimated dietary intake of PCBs is
approximately 0.008 pg/kg/day (average of the three most recent figures
reported in Table 7.2), the average daily intake via diet would be
0.56 pg (560 ng). This estimate indicates that consumption of food may
be a major source of PCB exposure in humans; however, the source of the
PCBs in food may be significant. In the recent years of the Total Diet
Study, the primary source of PCBs in the diet has been in the food
category meat-fish-poultry (Gartrell et al. 1986a, 1985a,b). FDA
chemists have found that the source of the PCBs in the meat-fish-poultry
composite is almost always due to the fish component (Jelinek and
Corneliussen 1976). This suggests that persons consuming less than the
average amounts of fish will be exposed to lower quantities of PCBs.
7.2.4.2 Fish and precipitation
The U.S. Fish and Wildlife Service has analyzed whole fish samples
collected nationwide for PCB residues as part of the National Pesticide
Monitoring Program (Schmitt et al. 1985). Between 1980 and 1981.
315 fish were collected from 107 stations nationwide. PCB residues were
detected in 94% of all fish, with the geometric mean concentration of
all Aroclors (wet weight) found to be 0.53 pg/g. This concentration is
lower than previous monitoring in 1976 to 1977 and 1978 to 1979, which
found concentrations of 0.88 and 0.85 pg/g, respectively. It should be
noted that these fish analyses pertain to whole fish samples, which are
composites of both the edible and nonedible portions of the fish.
Therefore, the concentrations reported may not necessarily reflect the
actual human exposure that will occur from oral consumption. Composite
fish samples taken from major tributaries and embayments of Lake
Superior and Lake Huron in 1983 contained PCB levels of 600 to
72,000 ng/g on a lipid basis (Jaffe et al. 1985). Analysis of 62 samples
of commercial fish (primarily from Lake Ontario) collected in 1980 found
levels of 0.11 to 4.90 ppm (Ryan et al. 1984).
Based on available monitoring data from the literature, the
following PCB ranges (in ng/L) in rainwater appear to be typical at the
various locations (Eisenreich et al. 1981): urban (10 to 250), rural
(1 to 50), Great Lakes (10 to 150), marine (0.5 to 10), and remote (1 to
30). PCB levels of 0.160 to 1.0 ng/L have been detected in snow from the
Antarctic (Tanabe et al. 1983). A review of PCB monitoring of
precipitation is available (Mazurek and Simoneit 1985).
-------
92 Section 7
Table 7.2. Estimated dietary intake of PCBs for adults,
infants, and toddlers (Mg/kg/day)
Fiscal year
1981-1982
1980
1979
1978
1977
1976
Adult
0.003
0.008
0.014
0.027
0.016
T*
Infant
ND°
ND
ND
0.011
0.02S
T
Toddler
ND
ND
ND
0.099
0.030
ND
°ND not detected.
*T = trace.
Source: Gartrell et al. 198Sa,b,c and 1986a,b.
-------
Potential for Human Exposure 93
7.3 OCCUPATIONAL EXPOSURES
1C was estimated that approximately 12,000 U.S. workers were
potentially exposed to PCBs annually from 1970 to 1976 (NIOSH 1977a). At
present, however, PCBs are no longer manufactured or used industrially
in the United States. Therefore, occupational exposure to those workers
involved in producing PCBs or manufacturing products with PCBs should no
longer occur. The potential for occupational exposure still exists,
however, since PCB-containing transformers and capacitors remain in use.
Exposure may occur during repair or accidents of electrical equipment
containing PCBs (Wolff 1985). Occupational exposure may also occur
during waste site cleanup of PCB-containing waste sites.
7.4 POPULATIONS AT HIGH RISK
Several groups are at high risk from PCBs because of unusually high
exposures. Persons occupationally exposed to PCBs are at high risk.
Nursing infants may be exposed to high PCB concentrations in the breast
milk of lactating women (EPA 1985a), especially if the women consume
large amounts of contaminated fish. Levels found in breast milk are
discussed in Sect. 2.2.3.1.
Other subpopulations are at high risk from PCBs because they are
more sensitive to toxic effects of exposure. Embryos, fetuses, and
neonates are potentially susceptible because of physiological
differences from adults. They generally lack the hepatic microsomal
enzyme systems that facilitate detoxification and excretion of PCBs
(Calabrese and Sorenson 1977, Gillette 1967, Nyhan 1961). Breast-fed
infants have additional risk caused by a steroid excreted in human
breast milk, but not cow's milk, that inhibits glucuronyl transferase
activity and thus glucuronidation and excretion of PCBs (Calabrese and
Sorenson 1977, Gartner and Arias 1966). Children exposed to the
antibiotic novobiocin may also be at greater risk because novobiocin
noncompetetively inhibits glucuronyl transferase activity in vitro
(Lokietz et al. 1963, Calabrese and Sorenson 1977).
Other subpopulations that are potentially more sensitive to PCBs
include those with incompletely developed glucuronide conjugation
mechanisms, such as those with Gilbert's syndrome or Crigler and Najjar
syndrome (Lester and Schmid 1964, Calabrese and Sorenson 1977). Persons
with hepatic infections may have decreased glucuronide synthesis, making
them more sensitive because of their decreased capacity to detoxify and
excrete PCBs (Calabrese and Sorenson 1977).
The indoor air in seven public buildings (schools, offices) was
monitored in Minnesota during 1984 for Aroclors 1242, 12S4, and 1260
(Oatman and Roy 1986). The total mean Aroclor concentration in the
indoor air of the three buildings using PCB transformers was found to be
nearly twice as high as that in the air of the four buildings not using
PCB transformers (457 ± 223 s.d. vs 229 ± 106 s.d. ng/m3). It is also
noteworthy that the levels found in all the indoor airs were
significantly higher than in typical ambient outdoor air.
The indoor air in a number of laboratories, offices, and homes was
monitored for various Aroclors. It was found that "normal* indoor air
concentrations of PCBs were at least one order of magnitude higher than
-------
94 Section 7
those in the surrounding outdoor atmosphere (MacLeod 1981). For example
average PCB levels were 0.10 pg/m3 inside an industrial research
building and 0.21 /*g/m3 inside the laboratories compared with
<0.02 /ig/m3 outside the facility. The average PCB level inside one home
was 0.31 Mg/mJ. while outside on the same day, the level was
0.004 pg/m3. It was suggested that certain electrical appliances and
devices (such as fluorescent lighting ballast). which have PCB-
containing components, can emit PCBs into the indoor air, thereby
elevating indoor PCB levels significantly above outdoor background
levels.
-------
95
8. ANALYTICAL METHODS
8.1 ENVIRONMENTAL MEDIA
The method widely used in laboratories for the analysis of PCBs in
complex environmental samples is capillary column gas chromatography
with electron capture (EC) detection (Schneider et al. 1984, Alford-
Stevens et al. 1986). The use of mass spectrometry (MS) detectors has
increased significantly, but most laboratories rely on EC detectors. EC
detectors are more sensitive than MS detectors operated in electron
ionization mode; the sensitivity difference can be as much as 2 or
3 orders of magnitude (Alford-Stevens et al. 1986). Table 8.1 lists
several analytical methods, which have been standardized by either the
EPA or NIOSH, for PCB analysis. The methods for water and for soil and
sediment that are required by the EPA Contract Laboratory Program (EPA
1987c) are designated as CLP on Table 8.1. Details of sample collection,
storage, and analysis of PCBs are available (Erickson 1986).
The analytical methods referenced in Table 8.1 pertain to the
detection of Aroclor formulations and not individual PCB isomers. With
EPA Method 680, however, PCBs are identified and measured by the level
of chlorination (EPA 1985c). This method has been used only since 1981,
and most environmental data reported before that were probably
underestimated.
The determination of Aroclor concentrations (rather than the level
of chlorination) in environmental samples is complex and can produce
significantly different results from different laboratories even though
the analytical procedures have been standardized (Alford-Stevens et al.
1985). As a result of the difference in biodegradability, water
solubility, and volatility of individual PCB isomers, the concentrations
of these individual isomers in environmental samples can be strikingly
different from the commercial PCB analytical reference standards.
8.2 BIOMEDICAL SAMPLES
Analytical methods used for biomedical samples are listed in Table
8.2. Gas chromatography-mass spectrometry procedures developed to
determine milligram-per-kilogram levels of PCBs in breast milk and fat
(Hutzinger et al. 1974) usually have lower sensitivity than EC detectors
(Safe et al. 1985, Smrek and Needham 1982). No accepted quantitative
procedure for the determination of the total PCB content in human tissue
sample exists. The PCB standard mixture selected for quantification
varies between investigators since no standard mixture exists with the
same peak pattern as in human tissues because of differences in
metabolism of the various PCB isomers. In recent years, high-resolution
gas chromatography has made it possible to use single PCB congeners for
quantitation. The selection of the congeners may be made on the basis of
-------
O\
IMC 0.1. HBBIJUOH (!
(Gamut* mtfni
Air
Air
Water
Water
Air
Soil. ttitinKiHi.
til nfluv mild
ano oucr souu
sample matrices
SoU/aodinent
(low level)
fiaiririff prcparatioo
Adsorption on (Ian filter and
Florisil; "MM* {tttflfirtifffl
Adsorption on Florinl; hcune
Extraction with methylene
chloride: dry eitract; exchange
loheiaoe
Eitnction with melbylene
riilnrSii^
duotioe
Adiorptioo on water-deactivated
Floriiil. beune deaorptian; per-
cblorination with antimony penta-
chloride at M8°C
Extraction with heune-aeetooc mixture.
up and dctuliuhzation by copper or
mercury, if necettary
Sample mixed with anhydrous sodium
sulfatc extracted with 1:1 melhylene
fhhifittff/afttHMtf conoentralc and
clean-up by gel permeation and
micro alumina column
Analytical method
MMB.Jm.ww. ..MrVMWW
GC/EC
CC/EC
GC/EC
GC/MS
GC/EC
GC/EC
GC/EC
Detection limit
0.0006 mg/ra'
forJO-L
sample
0.01 mg/rn'
(32 pg/iojectioo)
OOnSiig/L
(PCB-1242)
30-36 jig/L
(PCB-I22I. I2S4)
NR
-------
Sample malm Sample preparation
Blood strum Extract icnim with ethyl ether
Analytical method" Detection limit Accuracy/precision
HRGC/EC 1 0 ng/mL on 10-mL >80% accuracy at 25-400
References
NIOSH I984b
Tittue, eggs.
fat
Serum
Serum
Serum
Adipose
luiue
Human milk
Serum
Blood
and n-bexanc; treat with
melhanolic KOH; extract with
hexane and column chromaiographic
cleanup by silica gel
See Buih and Lo 1973 TLC
Mixed solvent extraction. GC/EC
column chromaiographic clean-
up on silica gel
0 S mg/kg
NR
NR
"8/mL (method 8004)
Precision ±005 mg/kg at (ARC 1978
0 5 mg/kg
Accuracy 92 6% at SO pg/L Burse el al I983a.b
and 1141% at lOjig/L.
accuracy 89 6-1381% at
9 9-74 2 pg/L for inter-
laboratory determinations
Accuracy 93 7% al 41 jig/L Needham el al 1980
2 S ng/mL
NR
Accuracy 95.3% al 100 »g/L Needham ct al 1981
and 105-127% at IOng/L
Accuracy 91-93% at 3 jig/g Smrek and Needham
1982
Solvent extraction, column GC/EC
chromaiographic cleanup on 10%
silver nitrate on silica gel
Mixed solvent extraction, column GC/EC
chromaiographic cleanup with
hydraled silica gel for
separation of PCBs from PBBs
Solvent extraction, column GC/EC
chromaiographic cleanup on
sulfuric acid/silica gel and
10% silver nitrate/silica
gel columns
Mixed solvent extraction. HRGC/EC
cleanup on Floruit-silicic
acid column
Solvent extraction with dieihyl HRGC/EC
ether and hexane. sulfunc acid.
and silica column cleanup
Solvent extraction with hexanc, GC/EC
melhanolic KOH hydrolysis, silica
gel. and alumina column cleanup
and perchlormalion
"HRGC - high-resolution gas chromalography. GC = gas chromatography. bC = electron capture. TLC = thin-layer chromaiography, NR = not
reported
NR
0.1 ng/mL
NR
NR
Mes el al 1984
85% at 25-125 ng/mL Luolamo el al 1985
NR
Lin and Que Hee 1985.
1987
n
i-
O
o
a
(A
vO
-------
98 Section 8
their abundance in the samples, their toxicity, or their availability in
analytical standards. A congener-specific analysis of a commercial PCB
preparation and the PCB composition of a human milk sample have been
reported by Safe et al. (1985). Variables in sampling method may also
greatly influence results. For example, PCB levels in milk fat may
decrease during lactation and with maternal age, weight, and purity
(Jensen 1987). It has been shown by Lawton et al. (1985) that random
error, interlaboratory variations in procedure, and methods used for
reporting data can all have considerable impact on the reported PCB
levels in human tissues. Such effects, however, should not deter
investigators from using serum PCB data for assessing environmental
exposure to populations or for statistical correlations with clinical
parameters in epidemiological studies. Caution should be exercised when
comparing exposure estimates or health effect studies reported by
different investigators or when considering "the use of a specific serum
PCB tolerance limit as a basis for administration action" (Lawton et al.
1985).
-------
99
9. REGULATORY AND ADVISORY STATUS
9.1 INTERNATIONAL
No data were located in the available literature.
9.2 NATIONAL
9.2.1 Regulations
9.2.1.1 Air
AGENCY ADVISORY
OSHA Chlorodiphenyl (42% chlorine)-Skin
TWA--1.0 mg/m3 (PEL) (OSHA 1985)
Chlorodiphenyl (54% chlorine)-Skin
TWA--0.5 mg/m3 (PEL) (OSHA 1985)
9.2.1.2 Food
FDA temporary tolerances
Agency Standard Value (ppm) References
FDA Foods 0.2-3.0 EPA 1988a
FDA Packaging 10.0 EPA 1988a
9.2.1.3 Vater
PCBs are prohibited in any discharge from any PCB manufacturer
(EPA 1977).
PCBs are regulated under the Clean Water Act Effluent Guidelines
for the following industrial point sources: electroplating, steam
electric, asbestos manufacturing, timber products processing, metal
finishing, paving and roofing, paint formulating, ink formulating, gum
and wood, carbon black, and aluminum forming (EPA 1988c).
9.2.2 Advisory Guidance
9.2.2.1 Air
AGENCY ADVISORY
PCBs
NIOSH REL-TWA--1.0 jig/m3, the minimum reliable
detectable concentration (NIOSH 1977b)
-------
LOO Section 9
American Conference of
Government Industrial
Hygienists (ACGIH)
ACGIH
9.2.2.2 Water
AGENCY
EPA
Aroclor 1254
TLV-TWA--0.5 mg/m3 (ACGIH 1986)
Aroclor 1242
TLV-TWA--1 mg/m3 (ACGIH 1986)
ADVISORY
Ambient water quality criteria (AVQO--0.79 to
0.0079 ng/L for carcinogenicity at 10" ^ to 10"7
risk levels (EPA 1980ba)
National Academy
of Sciences (NAS)
EPA
9.2.2.3 Soil
AGENCY
EPA
Drinking water criteria (DWC)--O.S to 0.005
for carcinogenicity at 10'^ to 10'6 risk levels
(EPA 1988a)
Suggested no adverse response level (SNARL)--
350 Mg/L (NAS 1980)
Aroclor 1016
Longer-term health advisory (HA) (adult)--
0.0035 mg/L (EPA 1988a)
Longer-term HA (child)--0.001 mg/L (EPA 1988a)
ADVISORY
Permissible PCB soil contamination levels corresponding to:
Noncancer 10-day HA (adult)--700 jig/day
Noncancer 10-day HA (child)--100 /ig/day
Cancer risk specific doses: 1.75 to 0.00175 jig/day at
10-* to 10-7 risk levels (EFA 1986d)
9.2.2.4 Other*
AGENCY ADVISORY
EPA Reportable quantity (RQ) (statutory)--10 Ib (EPA 1985d)
RQ (proposed)--1 Ib (EPA 1987d)
9.2.3 Data Analysis
Carcinogenic potency. EPA (1988a,b) determined that the positive
evidence for carcinogenicity of Aroclor 1254, Aroclor 1260, Kaneclor
500, and Clophen A-30 and A-60 in animals, along with inadequate
evidence in humans, places these PCBs in category B2, probable human
carcinogens. Because any PCB mixture that contains appreciable amounts
of the components in Aroclors 1254 and 1260, Kaneclor 500, and Clophen
A-30 and A-60 are likely to present a carcinogenic risk and because of
-------
Regulatory and Advisory Status 101
the variety and variability of PCB mixtures, EPA (1988a,b) recommended
that all commercial PCB mixtures be considered to have a similar
carcinogenic potential and classified all PCB mixtures in category B2.
IARC (1982) has classified PCBs in Group 2B based on sufficient evidence
in animals, inadequate evidence in humans, and inadequate evidence for
mutagenicicy. NIOSH (1986) recommended that PCBs be regarded as
potential human carcinogens in the workplace.
EPA (1988a,b) used the Norback and Weltman (1985) study as the
basis for a quantitative carcinogenicity risk assessment for PCBs. The
dietary level of 100 ppm Aroclor 1260 was converted to an intake of
5 mg/kg/day by assuming that a rat consumes food equal to 5% of its body
weight per day. This dosage was converted to a TWA dosage of
3.45 ing/kg/day to reflect the fact that rats received 100 ppm for
16 months, 50 ppm for 8 months, and 0 ppm for the last 5 months. The rat
dosage was converted to an equivalent human dose of 0.59 mg/kg/day on
the basis of relative body surface areas. Incidences of trabecular
carcinomas, adenocarcinomas, and neoplastic nodules in the liver were
combined to produce total incidences of 45/47 in treated females and
1/49 in controls. Using these data, EPA (1988a,b) calculated a human q *
of 7.7 (mg/kg/day)"1. Because there is no information regarding which
constituents of any PCB mixture might be carcinogenic, Aroclor 1260 is
assumed to be representative of other mixtures, and this potency
estimate applies to them as well (EPA 1988a,b). The q * was verified by
the EPA agency-wide CRAVE committee on April 22, 1987 (EPA 1988b).
9.3 STATE
Regulations and advisory guidance from the states were not
available.
-------
103
10. REFERENCES
ACGIH (American Conference of Governmental Industrial Hygienists). 1986.
Threshold Limit Values and Biological Exposure Indices for 1986-1987.
Cincinnati, Ohio.
Albro PW, Fishbein L. 1972. Intestinal absorption of polychlorinated
biphenyls in rats. Bull Environ Contain Toxicol 8:26 (cited in EPA
1985a).
Alford-Stevens AL, Bellar TA, Eichelberger JV, Budde WL. 1986. Accuracy
and precision of determinations of chlorinated pesticides and
polychlorinated biphenyls with automated interpretation of mass
spectrometric data. Anal Chem 58(9):2022-2029.
Alford-Stevens AL, Budde WL, Bellar TA. 1985. Interlaboratory study on
determination of polychlorinated biphenyls in environmentally
contaminated sediments. Anal Chem 57:2452-2457.
Allen JR. 1975. Response of the nonhuman primate to polychlorinated
biphenyl exposure. Fed Proc 34:1675-1679.
Allen JR, Abrahamson LJ. 1973. Morphological and biochemical changes in
the liver of rats fed polychlorinated biphenyls. Arch Environ Contain
Toxicol 1:265-280 (cited in EPA 1988a).
* Allen JR, Barsotti DA. 1976. The effects of transplacental and mammary
movement of the PCBs on infant rhesus monkeys. Toxicology 6:331.
* Allen JR, Barsotei DA, Carstens LA. 1980. Residual effects of
polychlorinated biphenyls on adult nonhuman primates and their
offspring. J Toxicol Environ Health 6(l):55-66 (cited in EPA 1988a).
* Allen JR, Barsotti DA, Lambrecht LK, Van Miller JP. 1979. Reproductive
effects of halogenated aromatic hydrocarbons on nonhuman primates. Ann
NY Acad Sci 320:419.
Allen JR, Carstens LA, Abrahamson LJ, Marlar RJ. 1975. Responses of rats
and nonhuman primates to 2,5,2',5'-tetrachlorobiphenyl. Environ Res
9:265-273 (cited in EPA 1988a).
* Key studies.
** No other names provided.
-------
104 Section 10
Allen JR, Carstens LA, Barsotti DA. 1974a. Residual effects of short-
term, low-level exposure of nonhuman primates to polychlorinated
biphenyls. Toxicol Appl Pharmacol 30:440-451.
Allen JR, Norback DH. 1973. Polychlorinated biphenyl and triphenyl
induced gastric mucosal hyperplasia in primates. Science 179:498.
Allen JR, Norback DH, Hsu 1C. 1974b. Tissue modifications in monkeys as
related to absorption distribution and excretion of polychlorinated
biphenyls. Arch Environ Contain Toxicol 2(l):86-95.
Alvares AP, Fischbein A, Anderson KE, Kappas A. 1977. Alterations in
drug metabolism in workers exposed to polychlorinated biphenyls. Clin
Pharmacol Ther 22:140.
Alvares AP, Kappas A. 1979. Lead and polychlorinated biphenyls: Effects
on heme and drug metabolism. Drug Hetab Rev 10:91-106.
Amano M, Yagi K, Nakajima H, Takehara R. Sakai H, Umeda G. 1984.
Statistical observations about the causes of death of patients with oil
poisoning. Japan Hygiene 39:1-5 (cited in EPA 1988a).
Anderson HA. 1985. Utilization of adipose tissue biopsy in
characterizing human halogenated hydrocarbon exposure. Environ Health
Perspect 60:127-131.
Anderson LM, Van Havere K, Budinger JM. 1983. Effects of polychlorinated
biphenyls on lung and liver tumors initiated in suckling mice by
ANnitrosodimethylamine. J Nat Cancer Inst 71(1):157-163 (cited in EPA
1985a).
Ando M, Saito H, Uakisaka I. 1985. Transfer of polychlorinated biphenyls
to newborn infants through the placenta and mothers' milk. Arch Environ
Contain Toxicol 14(1): 51-57.
* Aulerich RJ, Ringer RK. 1977. Current status of PCB toxicity,
including reproduction to mink. Arch Environ Contain Toxicol 6:279.
Bahn AK, Grover P. Rosenvaike I, O'Leary K, Stollman J. 1977. PCB and
melanoma. N Engl J Med 296:108 (cited in EPA 1988a).
Bahn AK, Rosenvaike I, Herrmann N, Grover P, Stollman J, O'Leary K.
1976. Melanoma after exposure to PCBs. N Engl J Med 295:450.
Baker EL, Landrigan PJ. Glueck CJ, et al. 1980. Metabolic consequences
of exposure Co polychlorinated biphenyls (PCB) in sewage sludge. Am J
Epidemiol 112:553-563.
Baker FD, Bush B, Tumasonls CF, Lo FC. 1977. Toxicity and persistence of
low-level PCBs in adult Vistar rats, fetuses, and young. Arch Environ
Contam Toxicol 5(2):143-156.
-------
References 105
Baker JE, Eisenreich SJ, Johnson TC, Halfman BN. 1985. Chlorinated
hydrocarbon cycling in the benthic nepreloid layer of Lake Superior.
Environ Sci Technol 19:854-861.
Bannister R, Davis D, Zacharewski T, Tizard I, Safe S. 1987. Aroclor
1254 as a 2,3,7,8-tetrachlorodibenzo-p-dioxin antagonist: Effects on
enzyme induction and inmunotoxicity. Toxicology (in press).
Barnes P, Bellin J, DeRosa C, et al. 1987. Reference dose (RfD):
Description and use in health risk assessments. Appendix A of the
Integrated Risk Information System. EPA 600/8-86-0321. Washington, D.C.:
Office of Health and Environmental Assessment, Office of Research and
Development.
* Barsotti DA, Allen JR. 1975. Effects of polychlorinated biphenyls on
reproduction in the primate. Fed Proc 34:338.
* Barsotti DA, Van Miller JP. 1984. Accumulation of a commercial
polychlorinated biphenyl mixture (Aroclor 1016) in adult rhesus monkeys
and their nursing infants. Toxicology 30(l):31-44.
* Barsotti DA, Marlar RJ, Allen JR. 1976. Reproductive dysfunction in
rhesus monkeys exposed to low levels of polychlorinated biphenyls
(Aroclor 1248). Food Cosmet Toxicol 14:99-103.
Becker GM, McNulty WP, Bell M. 1979. Polychlorinated biphenyls-induced
morphologic changes in the gastric mucosa of the rhesus monkey. Invest
40:373.
Bell M. 1983. Intrastructural features of the murine cutaneous
microvasculature after exposure to polychlorinated biphenyls compounds
(PCBs) and benzo(a)pyrene (BAP). Virchows Arch B 42(2):131-142 (cited in
EPA 1988a).
Benthe HF, Knop J, Schmoldt A. 1972. Absorption and distribution of
polychlorinated biphenyls (PCB) after inhalatory application. Arch
Toxicol 29:85.
Berry DL, DiGiovanni J, Juchau MR, Bracken WM, Gleason GL, Slaga TJ.
1978. Lack of tumor-promoting ability of certain environmental chemicals
in a two-stage mouse skin tumorigenesis assay. Res Commun Chem Pathol
Pharmacol 20(1):101-108.
Berry DL, Slaga TJ, DiGiovanni J, Juchau MR. 1979. Studies with
chlorinated dibenzo-p-dioxins, polybrominated biphenyls, and
polychlorinated biphenyls in a two-stage system of mouse skin
tumorigenesis: Potent anticarcinogenic effects. Ann NY Acad Sci
320:405-414.
Bertazzi PA, Riboldi L, Pesatori A, Radice L. Zocchetti C. 1987. Cancer
mortality of capacitor manufacturing workers. Am J Ind Med 11:165-176.
-------
106 Section 10
Bidleman TF. 1981. Interlaboracory analysis of high molecular weight
organochlorines in ambient air. Atmos Environ 15:619-624.
Billings RE, McMahon RE. 1978. Microsomal biphenyl hydroxylation: The
formation of 3 hydroxybiphenyl and biphenyl catechol. Mol Fharmacol
14:145-154.
Biocca M, Gupta BNL, Chae K, McKinney JD, Moore JA. 1981. Toxicity of
selected symmetrical hexachlorobiphenyl isomers in the mouse. Toxicol
Appl Fharmacol 58:461-474 (cited in EPA 1988a).
Birnbaum LS. Weber H, Harris MW, Lamb JC, McKinney JD. 1985. Toxic
interaction of specific polychlorinated biphenyls and 2,3,7,8-
tetrachlorodibenzo-p-dioxin: Increased incidence of cleft palate In
mice. Toxicol Appl Pharmacol 77:292-302.
Blazak WF, Marcun JB. 1975. Attempt to introduce chromosomal breakage in
chicken embryos with Aroclor 1242. Poultry Sci 54:310 (cited in Harbison
1986).
* Bleavins MR, Aulerich RJ, Ringer RK. 1980. Polychlorinated biphenyls
(Aroclors 1016 and 1242): Effects on survival and reproduction in mink
and ferrets. Arch Environ Contam Toxicol 9(5):627-635.
Bleavins MR, Breslin VJ. Aulerich RJ, Ringer RK. 1984. Placental and
mammary transfer of a polychlorinated biphenyl mixture (Aroclor 1254) in
the European ferret (Mustela pucorius furo). Environ Toxicol Chem
3(4):637-644.
Boon JP, Duinker JC. 1986. Monitoring of cyclic organochlorines in the
marine environments. Environ Monit Assess 7:189-208.
Bopp RF, Simpson HJ, Olsen CR, Trier RM, Kostyk N. 1982. Chlorinated
hydrocarbons and radionuclide chronologies in sediments of the Hudson
River and Estuary, N.Y. Environ Sci Techno 1 16:666.
Brezner E. Terkel J, Perry AS. 1984. The effect of Aroclor 1254 (PCB) on
the physiology of reproduction in the female rat--I. Comp Biochem
Physio1 77(1):65-70.
Brown DP. 1986. Mortality of Workers Exposed to Polychlorinated
Biphenyls --An Update. Cincinnati, Ohio: Industry Wide Studies Branch,
Div. of Surveillance, Hazard Evaluation and Field Studies, National
Institute National Institute for Occupational Safety and Health, Centers
for Disease Control, U.S. Public Health Service, Dept. of Health and
Human Services. NTIS PB86-206000.
Brown DP, Jones M. 1981. Mortality and industrial hygiene study of
workers exposed to polychlorinated biphenyls. Arch Environ Health
36(3):120-129.
-------
References 107
Brown JF, Jr., Bedard BL, Brennan MJ, Carnahan JC, Feng H, Wagner RE.
1987. Polychlorinated biphenyl dechlorination in aquatic sediments.
Science 236:709-712.
Brown JF, Jr., Lawton RW. 1984. Polychlorinated biphenyl (PCB)
partitioning between adipose tissue and serum. Bull Environ Contam
Toxicol 33:277-280.
Brown WR, Heddle JA. 1979. The mutagenic activity of 61 agents as
determined by the micronucleus, Salmonella, and sperm abnormality
assays. Can J Genet Cytol 21:319-333.
* Bruckner JV, Khanna KL, Cornish HH. 1973. Biological responses of the
rat to polychlorinated biphenyls. Toxicol Appl Pharmacol 24:434-448.
* Bruckner JV, Khanna KL, Cornish HH. 1974. Effect of prolonged
ingestion of polychlorinated biphenyls on the rat. Food Cosmet Toxicol
12:323.
Burkhard LP, Armstrong DE, Andren AW. 1985. Henry's law constants for
the polychlorinated biphenyls. Environ Sci Technol 19:590-596.
Burse VW, Needham LL, Korver MP et al. 1983a. Gas-liquid chromatographic
determination of polychlorinated biphenyls and a selected number of
chlorinated hydrocarbons in serum. J Assoc Off Anal Chem 66:32-39.
Burse VW, Needham LL, Lapeza CR, Jr., et al. 1983b. Evaluation of
potential analytical approach for determination of polychlorinated
biphenyls in serum: Interlaboratory study. J Assoc Off Anal Chem
66:956-968.
Bush B, Snow J, Koblintz. 1984. Polychlorobiphenyl (PCB) congeners,
p,p'-DDE, and hexachlorobenzene in maternal and fetal cord blood from
mothers in upstate New York. Arch Environ Contam Toxicol 13:517-527.
Calabrese EJ, Sorenson AJ. 1977. The health effects of PCBs with
particular emphasis on human high risk groups. Rev Environ Health
2(4):285-304 (cited in EPA 1988a).
Calandra JC. 1976. Summary of toxicological studies on commercial PCBs.
In: Proceedings of the National Conference of Polychlorinated Biphenyls.
EPA Report 560/6-75-004 (cited in Harbison 1986).
Callahan MA. Slimak MW, Gabel NW, et al. 1979. Water-related
environmental fata of 129 Priority Pollutants Vol. I. Chap 36. EPA
440/4-79-029a. Washington, DC: Environmental Protection Agency.
Carey AE, Gowen JA, Tai H, Mitchell WG. Wiersma GB. 1979a. Pesticide
residue levels in soils and crops from 37 states, 1972 - National Soils
Monitoring Program (IV). Pestic Monit J 12:209-229.
-------
108 Section 10
Carey AE, Douglas P, Tal H, Mitchell WG, Wlersma GB. 1979b. Pesticide
residue concentrations in soils of five United States cities, 1971 -
Urban Soils Monitoring Program. Pestle Monitor J 13:17-22.
* Carter JW. 1985. Effects of dietary PCBs (Aroclor 1254) on serum
levels of lipoprotein cholesterol in Fischer rats. Bull Environ Contain
Toxicol 34(3):427-431.
Chakraborty D, Bhattacharyya A, Chatterjee J, et al. 1978. Biochemical
studies on polychlorinated biphenyls toxicity in rats: Manipulation by
Vitamin C. Int J Vitam Nutr Res 48:22 (cited in EPA 1985a).
Chase KH, Wong 0, Thomas D, Berney BV, Simon RK. 1982. Clinical and
metabolic exposure to polychlorinated biphenyls (PCBs). J Occup Med
24:109-114 (cited in Kreiss 1985).
Chemline. 1987. On-line computer data base. National Library of
Medicine. June 4, 1987.
Chen PH, Luo ML, Wong CK, Chen CJ. 1982. Comparative rates of
elimination of some individual polychlorinated biphenyls from the blood
of PCB-poisoned patients in Taiwan. Food Chem Toxicol 20(4):417-425.
Chen PH, Wong CK, Rappe C, Nygren M. 1985. Polychlorinated biphenyls,
dibenzofurans and quaterphenyls in toxic rice-bran oil and in the blood
and tissues of patients with PCB poisoning (Yu-Cheng) in Taiwan. Environ
Health Perspect 29:475-678 (cited In EPA 1988a).
Christensen ER. Lo CK. 1986. Polychlorinated biphenyls in dated
sediments of Milwaukee Harbor, Wisconsin. Environ Pollut 12:217-232.
Chu CK. Stella VJ, Bruckner JV, Jiang WD. 1977. Effects of long-term
exposure to environmental levels of polychlorinated biphenyls on
pharmacokinetics of pentobarbital in rats. J Pharm Sci 66(2):238-241
(cited in EPA 1988a).
* Collins WT, Capen CC. 1980a. Fine structural lesions and hormonal
alterations in thyroid glands of perinatal rats exposed in utero and by
milk to polychlorinated biphenyls. Am J Pathol 99:125-142.
Collins WT, Capen CC. 1980b. Biliary excretion of thyroxine-I-125 and
fine structural alterations in the thryoid glands of gunn-rats fed PCBs.
Lab Invest 43:158.
Collins WT, Capen CC. 1980c. Ultrastructural and functional alterations
of the rat thyroid gland produced by polychlorinated biphenyls compared
with iodide excess and deficiency, and thyrotropln and thyroxlne
administration. Virchos Arch B: 33(3):213-231.
Collins WT, Capen CC, Kasza L, Carter C, Dailey RE. 1977. Effect of
polychlorinated biphenyl (PCB) on the thyroid gland of rats.
Ultrastructural and biochemical investigations. Am J Pathol 89:119.
-------
References 109
Colombi A, Maroni M, Ferioli A et al. 1982. Increase in urinary
porphyrin excretion in workers exposed to polychlorinated biphenyls. J
Appl Toxicol 2(3):117-121.
Condon SK. 1983. (Commonwealth of Massachusetts Department of Public
Health). Personal Communications, August 25 and 28, 1983 (cited in
Kreiss 1985).
Conolly RB, Szabo S, Jaeger RJ. 1979. Vinylidene fluoride. Acute
hepatotoxicity in rats pretreated with PCB or phenobarbital. Proc Exp
Biol Med 162:163 (cited in EPA 1985a).
Greaser CS, Fernandes AR. 1986. Background levels of polychlorinated
biphenyls in British soils. Chemosphere 15:499-508.
Davidorf FH, Knupp JA. 1979. Epidemiology of ocular melanoma. Incidence
and geographic relationship in Ohio (1967-1977). Ohio State Med J
75(9):561-564.
DiGiovanni J, Viaje A, Berry DL, Slaga TJ, Juchau MR. 1977. Tumor-
initiating ability of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and
Aroclor 1254 in the two-stage system of mouse skin carcinogenesis. Bull
Environ Contam Toxicol 18(5):552-557.
Dikshith TSS, Rockwood W, Abraham R, Coulston F. 1975. Effects of
polychlorinated biphenyls (Aroclor 1254) on rat testis. Exp Mol Pathol
22:376 (cited in EPA 1988a).
Drill VA, Freiss SL, Hays HW, Loomis TA, Shaffer CB. 1981. Potential
Health Effects in the Human from Exposure to Polychlorinated Biphenyls
(PCBs) and Related Impurities. Unpublished report. Arlington, Va: Drill,
Freiss, Hays, Loomis and Shaffer, Inc.
Drotman DP, et al.** 1981. Human Exposure to PCBs in Southern Idaho.
Internal report. EPA-79-105-2, Atlanta: Centers for Disease Control,
November 2, 1981 (cited in Kreiss 1985).
Drotman DP, Baxter PJ, Liddle JA, Brokopp CD, Skinner MD. 1983.
Contamination of the food chain by polychlorinated biphenyls from a
broken transformer. Am J Public Health 73:290-292.
Duggan RE, Corneliussen PE, Duggan MB, McMahon BM, Martin RJ. 1983.
Pesticide Residue Levels in Foods in the United States from July 1,
1969, to June 30, 1976. Washington, D.C.: Food and Drug Administration,
Division of Chemical Technology.
Durfee RL. 1976. Production and usage of PCBs in the United States. In:
Proceedings of the National Conference on Polychlorinated Biphenyls,
Chicago, 1975. EPA-560/6-75-004. Washington, D.C.: Environmental
Protection Agency, pp. 103-107.
-------
110 Section 10
Eduljee G, Badsha K, Price L. 1985. Environmental monitoring for PCB and
heavy metals in the vicinity of a chemical waste disposal facility-I
Chemosphere 14:1371-1382. ="«* i.
Eduljee G, Badsha K, Scudamore N. 1986. Environmental monitoring for PCB
and trace metals in the vicinity of a chemical waste disposal facilitv-
II. Chemospere 15:81-93.
Eisenreich SJ, Looney BB, Thornton JD. 1981. Airborne organic
contaminants in the Great Lakes ecosystem. Environ Sci Technol 15:30-38.
Emmett EA. 1985. Polychlorinated biphenyl exposure and effects in
transformer repair workers. Environ Health Perspect 60:185-192.
EPA (Environmental Protection Agency). 1976. PCBs in the United States.
Industrial Use and Environmental Distribution. PB-252 012. Springfield,
Va: National Technical Information Service, pp. 4-5. 34-35 54-57 198-
210, 322-334 (cited in IARC 1978).
EPA (Environmental Protection Agency). 1977. Polychlorinated biphenyls
(PCBs): Toxic pollutant effluent standards, final rule. Fed Resist
42(22):6531-6555.
EPA (Environmental Protection Agency). 1979. Polychlorinated biphenyls
(PCBs): Proposed rulemaking for PCB manufacturing exemptions. Fed Resist
44(106):31564-31567. *
EPA (Environmental Protection Agency). 1980a. Hazard Waste Generation
and Commercial Hazardous Waste Management Capacity: An Assessment.
SW-894. Washington. D.C.: EPA, p. D-4.
EPA (Environmental Protection Agency). 1980b. Ambient Water Quality
Criteria for Polychlorinated Biphenyls. EPA 440/5-80-068. Washington
D.C.: EPA. NTIS PB81-117798.
EPA (Environmental Protection Agency). 1982a. Test Methods. Methods for
Organic Chemical Analysis of Municipal and Industrial Wastewater. EPA
600/4-82-057. Cincinnati. Ohio: EPA. pp. 608-1 - 608-11; 625-1 - 625-12.
EPA (Environmental Protection Agency). 1982b. Test Method* for
Evaluating Solid Waste. SW-846. Washington, D.C.: Office of Solid Waste
and Emergency Response, EPA. pp. 8080-1 - 8080-17.
EPA (Environmental Protection Agency). 1985*. Drinking Water Criteria
Document for Polychlorinated Biphenyls (PCBs). Draft. Washington. DC:
Office of Drinking Water. NTIS PB 86-118312/AS.
EPA (Environmental Protection Agency). 1985b. Health Assessment Document
for Polychlorinated Dibenzo-p-Dioxins. EPA/600/8-84/014F, pp. II-1 -
11-29; IV-1 - IV-37.
-------
References 111
EPA (Environmental Protection Agency). 1985c. Method 680, Determination
of Pesticides and PCBs in Water and Soil/Sediment by Gas
Chromatography/Mass Spectrometry. Cincinnati, Ohio: Environmental
Monitoring and Support Laboratory, Office of Research and Development,
Environmental Protection Agency (cited in Alford-Stevens 1986).
EPA (Environmental Protection Agency). 1985d. Notification requirements,
reportable quantity adjustments, final rule and proposed rule. Fed
Regist 50(65):13456-13523.
EPA (Environmental Protection Agency). 1985e. Baseline Estimates and
Time Trends for Beta-Benzene Hexachloride, Hexachlorobenzene, and
Polychlorinated Biphenyls in Human Adipose Tissue 1970-1983. EPA 560/5-
85-025. Washington. D.C.: Office of Toxic Substances, Exposure
Evaluation Division. Doc. No. NHATS-SS-01.
EPA (Environmental Protection Agency). 1986a. Reference Values for Risk
Assessment. Prepared by the Office of Health and Environmental
Assessment for the Office of Solid Waste. Washington, D.C. Cincinnati,
Ohio: Environmental Criteria and Assessment Office.
EPA (Environmental Protection Agency). 1986b. Broad scan analysis of the
FY 82 national human adipose tissue survey specimens. Volume III -
Semi-Volatile Organic Compounds. EPA-560/5-86-037. Washington, D.C.:
Office of Toxic Substances.
EPA (Environmental Protection Agency). 1986c. Guidelines for carcinogen
risk assessment. Fed Regist 51(185):33992-34003.
EPA (Environmental Protection Agency). 1986d. Development of Advisory
Levels for Polychlorinated Biphenyls (PCBs) Cleanup. EPA/600/6-86-02.
Washington, D.C.
EPA (Environmental Protection Agency). 1987a. Polychlorinated biphenyl
spills cleanup policy; final rule. Fed Regist 52(63):10688-10710.
EPA (Environmental Protection Agency). 1987b. Graphical Exposure
Modeling System (GEMS). Personal computer version April 1987. Research
Triangle Park, N.C.: EPA.
EPA (Environmental Protection Agency). 1987c. U.S. Contract Laboratory
Program. Statement of Work for Organics Analyses, Multi-Media. Multi-
Concentration. Revised 8/87.
EPA (Environmental Protection Agency). 1987d. Reportable quantity
adjustments. Proposed rule. Fed Regist 52(50):8140.
EPA (Environmental Protection Agency). 1988a. Drinking Water Criteria
Document for Polychlorinated Biphenyls (PCBs). ECAO-C1N-414. Final.
April 1988.
-------
112 Section 10
EPA (Environmental Protection Agency). 1988b. IRIS (Integrated Risk
Information System), CRAVE (Carcinogen Risk Assessment Validation
Endeavor) for polychlorinated biphenyls. (Verification date: 4/22/87).
On-line: input pending. Cincinnati, Ohio: Office of Health and
Environmental Assessment, Environmental Criteria and Assessment Office.
EPA (Environmental Protection Agency). 1988c. Analysis of Clean Water
Act Effluent Guidelines Pollutants. Summary of the Chemicals Regulated
by Industrial Point Source Category, 40 CFR Parts 400-475. Draft.
Prepared by Industrial Technology Division (WH 552). Washington, D.C.:
Office of Water Regulations and Standards, Office of Water,
Environmental Protection Agency.
EPA-NIH (Environmental Protection Agency-National Institutes of Health).
1987. OHM-TADS (Oil and Hazardous Materials Technical Assistance Data
System). On-line: 1987. Washington, DC. EPA-NIH.
Erickson, MD. 1986. Analytical chemistry of PCBs. Stoneham, Mass.:
Butterworth Publishers, pp. 55-338.
Fein GG. 1984. Intrauterine exposure of humans to PCBs: Newborn effects.
EPA-600/53-84-060. Duluth, Minn: EPA. PB-84-188-887.
Fein GG, Jacobson JL, Jacobson SW, Schwartz PM, Dowler JK. 1984.
Prenatal exposure to polychlorinated byphenyls - effects on birth size
and gestational age. J Pediatrics 105:315-320.
Felt GR, Mueller WF, latropoulos MJ, Coulston F, Korte F. 1977. Chronic
toxicity of 2,5,4'-trichlorobiphenyl in young rhesus monkeys. I. Body
distribution elimination and metabolism. Toxicol Appl Pharmacol
41(3):619-627 (cited in EPA 1988a).
Feltz HR. 1980. Significance of bottom material data in evaluation water
quality. In: Contain Sed Fate Transport Case Studies Model Tox. Ann
Arbor, Mich: Ann Arbor Science 1:271-287.
Finklea J, Priester LE, Creason JP, Hauser T, Hinners T, Hammer D. 1972.
I. Polychlorinated biphenyl residues in human plasma expose a major
urban pollution problem. Am J Public Health 62:645-651 (cited in Kreiss
1985).
Fischbein A. 1985. Liver function tests in workers with occupational
exposure to polychlorinated biphenyls (PCBs): Comparison with Yusho and
Yu-Cheng. Environ Health Perspect 60:145-150.
Fischbein A, Rizzo JN, Solomon SJ, Wolff MS. 1985. Oculodermatological
findings in workers with occupational exposure to polychlorinated
biphenyls. Br J Ind Med 42(6):426-430.
Fischbein A, Wolff MS, Bernstein. Selikoff IJ. 1982. Dermatological
findings in capacitor manufacturing workers exposed to dielectric fluids
containing polychlorinated biphenyls. Arch Environ Health 37:69-74.
-------
References 113
* Fischbein A, Wolff MS, Lills R, Thornton J, Selikoff IJ. 1979.
Clinical findings among PCB-exposed capacitor manufacturing workers Ann
NY Acad Sci 320:703-715.
* Fishbein L. 1974. Toxicity of chlorinated biphenyls. Ann Rev Pharmacol
14:139-156.
Frank R. Braun HE, Van Hoveholdrinet M, Sirons GJ, Ripley BD. 1982.
Agriculture and water quality in the Canadian Great Lakes Basin: V.
Pesticide use in 11 agricultural watersheds and presence in steam water
1975-77. J Environ Qual 11:497.
Gage JC, Holm S. 1976. The influence of molecular structure on the
retention and excretion of polychlorinated biphenyls by the mouse
Toxicol Appl Pharmacol 36:555-560.
* Garthoff LH, Cerra FE, Marks EM. 1981. Blood chemistry alteration in
rats after single and multiple gavage administration of polychlorinated
biphenyls. Toxicol Appl Pharmacol 60(l):33-44.
Garthoff LH, Friedman L, Farber TM, et al. 1977. Biochemical and
cytogenetic effects in rats caused by short-term ingestion of Aroclor
1254 or Firemaster BP6. J Toxicol Environ Health 3:769 (cited in EPA
1988a).
Gartner LM, Arias IM. 1966. Studies of prolonged neonatal jaundice in
the breast-fed infant. J Pediat 68(1):54 (cited in EPA 1988a).
Gartrell MJ, Craun JC. Podrebarac DS, Gunderson EL. 1985a. Pesticides,
selected elements, and other chemicals in adult total diet samples
October 1979 - September 1980. J Assoc Off Anal Chem 68:1184-1197.
Gartrell MJ, Craun JC, Podrebarac DS, Gunderson EL. 1985b. Pesticides,
selected elements, and other chemicals in adult total diet samples
October 1979 - September 1980. J Assoc Off Anal Chem 68:862-873.
Gartrell MJ, Craun JC, Podrebarac DS, Gunderson EL. 1985c. Pesticides,
selected elements, and other chemicals in infant and toddler diet
samples, October 1979 - September 1980. J Assoc Off Anal Chem 68:1163-
1183.
Gartrell MJ, Craun JC, Podrebarac DS, Gunderson EL. 1986a. Pesticides,
selected elements, and other chemicals in adult total diet samples
October 1980 - March 1982. J Assoc Off Anal Chem 69:146-161.
Gartrell MJ, Craun JC, Podrebarac DS, Gunderson EL. 1986b. Pesticides,
selected elements, and other chemicals in infant and toddler total diet
samples October 1980 - March 1982. J Assoc Off Anal Chem 69:123-145.
Giam CS, Chan HS, Neff GS, Atlas EL. 1978. Phthalate ester plasticizers:
A new class of marine pollutant. Science 199:419-421.
-------
114 Section 10
Gillette JR. 1967. Individually different responses to drugs according
to age, sex and functional or pathological state. In: Volstenhome G,
Proter R, eds. Drug Responses in Man. London: Churchill, p. 28 (cited in
EPA 1988a).
Goldstein JA, Hickman P, Jue DL. 1974. Experimental hepatic porphyria
induced by polychlorinated biphenyls. Toxicol Appl Pharmacol 27:437-448.
Goto M, Sugiura K, Hattori M, Miyagawa T, Okamura M. 1974. Metabolism of
2,3-dichlorobiphenyl-14C and 2,4,6-trichlorobiphenyl-14C in the rat.
Chemosphere 5:227-232 (cited in EPA 1988a).
Grant DL, Phillips WEJ. 1974. The effect of age and sex on the toxicity
of Aroclor 1254, a polychlorinated biphenyl, in the rat. Bull Contain
Toxicol 12:145-152 (cited in EPA 1988a).
Green S, Carr JV, Palmer KA, Oswald EJ. 1975a. Lack of cytogenetic
effects in bone marrow and spermatogonial cells in rats treated with
polychlorinated biphenyls (Aroclors 1242 and 1254). Bull Environ Contam
Toxicol 13:14-22.
Green S, Sauro EM, Friedman L. 1975b. Lack of dominant lethality in rats
treated with polychlorinated biphenyls (Aroclor 1242 and 1254). Food
Cosmet Toxicol 13:507-510.
Griffin RA, Chou SFJ. 1981. Movement of PCBs and other persistent
compounds through soil. Water Sci Technol 13:1153-1163.
Gustavsson P, Hogstedt C, Rappe C. 1986. Short-term mortality and cancer
incidence in capacitor manufacturing workers exposed to polychlorinated
biphenyls. Am J Ind Med 10:341-344.
Guzelian PS. 1985. Clinical evaluation of liver structure and function
in humans exposed to halogenated hydrocarbons. Environ Health Perspect
60:159-164.
* Haake JM. Safe S, Mayura K, Phillips TD. 1987. Aroclor 1254 as an
antogonist of the teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin.
Toxicol Lett (in press).
Hansch C, Leo AJ. 1985. Medchem Project. Issue No. 26. Claremont, Calif:
Pomona College.
Harbison RD. 1986. Genotoxic effects of PCBs. Draft report sent to
Dr. John Craddock. St. Louis, Mo: Monsanto Chem.
Hashimoto K, Akasaka S, Takagi Y, et al. 1976. Distribution and
excretion of [14C]polychlorinated biphenyls after their prolonged
administration to male rats. Toxicol Appl Pharmacol 37:415-423.
Hatton RE. 1979. Chlorinated biphenyls and related compounds. In:
Grayson M, Eckroth D, eds. Kirk-Othmer Encyclopedia of Chemical
Technology, Vol. 5, New York: John Viley and Sons, pp. 844-848.
-------
References 115
Hattula ML. 1985. Mutagenlcity of PCBs and their pyrosynthetic
derivatives in cell-mediated assay. Environ Health Perspect 60:255-257.
Hawley JK. 1985. Assessment of health risk for exposure to contaminated
soil. Risk Anal 5(4):289-302.
Meddle JA, Bruce WR. 1977. Comparison of tests for mutagenicity or
carcinogenicity using assays for sperm abnormalities, formation of
micronuclei and mutations in Salmonella. In: Hiatt HH, et al., ed.
Origins of Human Cancer. Cold Spring Harbor Conference on Cell
Proliferation. Cold Spring Harbor, N.Y.: Cold Spring Harbor Lab 4-1549
(cited in EPA 1988a).
Heit M, Klusek C, Baron J. 1984. Evidence of deposition of anthropogenic
pollutants in remote Rocky Mountain lakes. Water Air Soil Pollut
22:403-416.
Hill RH, Jr. 1985. Effects of polyhalogenated aromatic compounds on
porphyrin metabolism. Environ Health Perspect 60:139-143.
Hollifield HC. 1979. Rapid nephelometric estimate of water solubility of
highly insoluble organic chemicals of environmental interest. Bull
Environ Contam Toxicol 23:579-586.
Hoopingarner R, et al.** 1972. Polychlorinated biphenyl interactions
with tissue culture cells. Environ Health Perspect 1:155 (cited in
Harbison 1986).
* Hornshaw TC, Safronoff J, Ringer RK, Aulerich RJ. 1986. LCso test
results in polychlorinated biphenyl-fed mink: Age, season and diet
comparisons. Arch Environ Contam Toxicol 15(6):717-723.
HSDB (Hazardous Substances Data Bank). 1987. On-line computer data base.
National Library of Medicine. June 4, 1987.
Hubbard HL. 1964. Chlorinated biphenyl and related products. In: Standen
A, ed. Kirk-Othmer Encyclopedia of Chemical Technology, 2nd ed., Vol. 5,
New York: John Wiley and Sons, p. 291.
Humphrey HEB. 1976. Evaluation of changes of the level of
polychlorinated biphenyls (PCB) in human tissue. Final Report on FDA
Contract 223-73-2209. Lansing, Michigan: Michigan Department of Public
Health.
Humphrey HEB. 1983a. Population studies of PCBs in Michigan residents.
In: D'ltri FM, Kamrin MA, eds. PCBs: Human and Environmental Hazards.
Ann Arbor, Mich: Ann Arbor Science Publications, pp. 299-310 (cited in
Kreiss 1985).
Humphrey HEB. 1983b. Evaluation of humans exposed to waterborne
chemicals of the Great Lakes. Final report for EPA Co-operative
Agreement (CR807192) (cited in Kreiss 1985).
-------
116 Section 10
Hutcon JJ, Meier J, Hackney C. 1979. Comparison of the in vitro
mutagenicity and metabolism of dime thy Initrosamine and benzo[a]pyrene in
tissues from inbred mice treated with phenobarbitol, 3-
methylcholanthrene or polychlorinated biphenyls. Hutat Res 66:75 (cited
in EPA 1985a).
Hutzinger S, Safe S, Zitko V. 1974. The chemistry of PCBs. Cleveland,
Ohio: Chemical Rubber Publishing Co (cited in Callahan et al. 1979, IARC
1978).
latropoulos MJ, Bailey J, Adams HP, Coulston, Hobson V. 1978. Response
of nursing infant rhesus to Clophen A-30 or hexachlorobenzene given to
their lactating mothers. Environ Res 16(1-3):38-47 (cited in EPA 1988a).
IARC (International Agency for Research on Cancer). 1978. IARC
monographs on the evaluation of the carcinogenic risk of chemicals to
humans. Polychlorinated Biphenyls and Polybrominated Biphenyls. IARC,
Vol. 18. Lyon, France: World Health Organization.
IARC (International Agency for Research on Cancer). 1982. IARC
monographs on the evaluation of the carcinogenic risk of chemicals to
humans. Suppl 4. Lyon, France: World Health Organization.
Ito N, Nagasaki H, Makiura S, Aral M. 1974. Histopathological studies on
liver tumorigenesis in rats treated with polychlorinated biphenyls. Gann
66:545-549 (cited in EPA 1985a).
Jacobson JL, Fein GG, Jacobson SW, et al. 1984a. The transfer of
polychlorinated biphenyls (PCBs) and polybrominated biphenyls (PBBs)
across the human placenta and into maternal milk. Am J Public Health
74(4):378-379.
Jacobson JL, Jacobson SW, Schwartz PH, Fein GG, Dowler JK. 1984b.
Prenatal exposure to an environmental toxic: A test of the multiple
effects model. Dev Psych 20:523-532.
Jacobson SW, Fein GG, Jacobson JL, Schwartz PM, Dowler JK. 1985. The
effect of intrauterine PCB exposure on visual recognition memory. Child
Dev 56:856-860.
Jaffe R, Stemmler EA, Eitzer BD, Hites RA. 1985. Anthropogenic,
polyhalogenated, organic compounds in sedentary fish from Lake Huron and
Lake Superior tributaries and embayments. J Great Lakes Res 11:156-162.
Jelinek CF. Corneliussen PE. 1976. Levels of PCBs in the U.S. food
supply. In: Proceedings of the National Conference on Polychlorinated
Biphenyls, Chicago, 1975. EPA-560/6-75-004. Washington, D.C.:
Environmental Protection Agency, pp. 147-154.
Jensen AA. 1983. Chemical contaminants in human milk. Res Rev 89:1, 75.
82-94.
-------
References 117
Jensen AA. 1987. Polychlorobiphenyls (PCBs), polychlorodibenzo-p-dioxins
(FCDDs) and polychlorodibenzofurans (PCDFs) in human milk, blood, and
adipose tissue. Sci Total Environ 64:259-293.
Jensen RG, Clark RM, Ferris AM. 1980. Composition of the lipids in human
milk: A review. Lipids 15:345-355 (cited in Kimbrough 1987a)
Jensen S, Sundstrom G. 1974. Structures and levels of most
chlorobiphenyls in the technical PCB products and in human adipose
tissue. Ambio 3:70-76 (cited in EPA 1988a).
Kasza L, Collins WT, Capen CC, Garthoff LH, Friedman L. 1978.
Comparative toxicity of polychlorinated biphenyls and polybrominated
biphenyl in the rat thyroid gland: Light and electron microscopic
alterations after subacute dietary exposure. J Environ Pathol Toxicol,
May-June (5):587-599.
Kato N, Kawai K, Yoshida A. 1981. Effect of dietary level of ascorbic
acid on the growth, hepatic lipid peroxidacion, and serum lipids in
guinea pigs fed polychlorinated biphenyls, Aroclor 1254. Bull Environ
Contain Toxicol 18:243 (cited in EPA 1985a) .
Keplinger ML, et al.** 1971. Toxicological studies with polychlorinated
biphenyls. Toxicol Appl Pharmacol 53:389 (cited in Harbison 1986).
Kim NK, Stone DU. n.d. Organic chemicals and drinking water. NYS Dept
Health, p. 101.
Kimbrough RD. 1987a. Human health effect of polychlorinated biphenyls
(PCBs) and polybrominated biphenyls (PBBs). Ann Rev Pharmacol Toxicol
27:87.
Kimbrough RD. 1987b. Toxicology of halogenated biphenyls,
dibenzodioxins, and dibenzofurans. ISI Atlas of Sciences: Pharmacology
1:139-142.
Kimbrough RD, Linder RE. 1974. Induction of adenofibrosis and hepatomas
in the liver of Balb/CJ mice by polychlorinated biphenyls (Aroclor
1254). J Nat Cancer Inst 53:547 (cited in EPA 1988a).
Kimbrough RD, Linder RE, Gaines TB. 1972. Morphological changes in
livers of rats fed polychlorinated biphenyls. Arch Environ Health
25:354.
Kimbrough RD, Squire TA, Linder RE. Strandberg JD. Montali RJ, Burse W.
1975. Induction of liver tumors in Sherman strain female rats by
polychlorinated biphenyl Aroclor 1260. J Nat Cancer Inst 55:1453-1459.
Kleinert JJ. 1976. Sources of polychlorinated biphenyls in Wisconsin.
In: Proceedings of the National Conference on Polychlorinated Biphenyls,
Chicago, 1975. EPA-560/6-75-004. Washington, D.C.: Environmental
Protection Agency, pp. 124-126.
-------
118 Section 10
Kokoszka L. Flood J. 1985. A guide Co EPA-approved PCB disposal methods
Chem Eng 92(14):41-43.
Roller LD. 1977. Enhanced polychlorinated biphenyls lesions in Moloney
leukemia virus-infected mice. Clin Toxicol 11(1):107-116.
Kraul I, Karlog 0. 1976. Persistent organochlorinated compounds in human
organs collected in Denmark 1972-73. Acta Pharmacol Toxicol (Kbh)
38(2):38-73 (cited in EPA 1985a).
Kreiss K. 1985. Studies on populations exposed to polychlorinated
biphenyls. Environ Health Perspect 60:193-199.
Kreiss K, Roberts C, Humphrey HEB. 1982. Serial PBB levels, PCB levels,
and clinical chemistries in Michigan's PBB cohort. Arch Environ Health
37:141-147 (cited in Kreiss 1985).
Kreiss K, Zack MM, Kimbrough RD, Needham LL, Smrek AL, Jones BT. 1981.
Association of blood pressure and polychlorinated biphenyl levels. J Am
Med Assoc 245(24):2505-2509.
Kurachi M. 1983. A new sulfur-containing derivative and possibility of
conjugate formation of PCBs in mice or rats. Agric Biol Chem
47(6):1183-1191.
Kurachi M, Mio T. 1983a. On fluctuation of PCBs under various unnatural
conditions in mice. Agric Biol Chem 47(6):1173-1181.
Kurachi M, Mio T. 1983b. Studies on excretion and accumulation of PCBs
in connection with their partial metabolism in the animal body. Part
III. On the formation of a conjugate of PCBs with glutathione and its
further metabolism in mice or rats. Agric Biol Chem 47(6):1193-1199.
Kuratsune M. 1986. Letter to A Chiu and 0 Bayliss. Carcinogen Assessment
Group, Washington, D.C.: Environmental Protection Agency. June 30 (cited
in EPA 1988a).
Kuratsune M, Shapiro R. 1984. PCB poisoning in Japan and Taiwan. Am J
Ind Med 5:1-153.
Larsson P. 1985. Contaminated sediments of lakes and oceans act as
sources of chlorinated hydrocarbons for release to water and atmosphere.
Nature 317:347-349.
Lawrence C. 1977. PCB? and melanoma. New Engl J Med 296:108.
Lawton RW, Brown JF, Ross MR, Feingold J. 1982. Comparability and
precision of serum PCB measurements. Arch Environ Health 40:29-37.
Lawton RW, Ross MR, Feingold J, Brown JF. Jr. 1985. Effects of PCB
exposure on biochemical and hematological finding in capacitor workers.
Environ Health Perspect 60:165-184.
-------
References 119
Lelfer A, Brink RH, Thorn GC, Partymlller KG. 1983. Environmental
transport and transformation of polychlorinated biphenyls. EPA-560/5-
83-025. Washington, D.C.: Office of Pesticides and Toxic Substances
NTIS No. PB84-142579.
Lester R, Schmid R. 1964. Bilirubin metabolism. New Engl J Hed
270(15):779 (cited in EPA 1985a).
Letz G. 1983. The toxicology of PCB's - an overview for clinicians. The
Western J Med 138:534-540.
Lewis RG, Martin BE, Sgontz DL, Howes JE, Jr. 1985. Measurements of
fugitive atmospheric emissions of polychlorinated biphenyls from
hazardous waste landfills. Environ Sci Technol 19:986-991.
Lin JM. Que Hee SS. 1985. Optimization of perchlorination conditions for
some representative polychlorinated biphenyls. Anal Chem 57:2130-2134.
Lin JM, Que Hee SS. 1987. Change in chromatogram patterns after
volatilization of some Aroclors, and the associated quantitation
problems. Am Ind Hyg Assoc J 48:599-607.
* Linder RE, Gaines TB, Kimbrough RD. 1974. The effect of PCB on rat
reproduction. Food Cosmet Toxicol 12:63.
* Litterst CL, Farber TM, Baker AM, van Loon EJ. 1972. Effect of
polychlorinated biphenyls on hepatic microsomal enzymes in the rat.
Toxicol Appl Pharmacol 23:112-122.
Lokietz H, Dowben RM, Hsia DY. 1963. Studies on the effect of Novobiocin
and glucuronyl transferase. Pediatrics 32:47 (cited in EPA 1985a).
Loose LD, Pittman KA, Benitz KF, Silkworth JB, Mueller W, Coulston F.
1978a. Environmental chemical- induced immune dysfunction. Ecotoxicol
Environ Safety 2:173.
Loose LD, Silkworth JB, Pittman KA. Benitz KF. Mueller V,. 1978b.
Impaired host resistance to endotoxic and malaria in polychlorinated
biphenyl and hexachlorobenzene-treated mice. Inf Immun 20(1):30.
Lucas RM, lannacchlone VG, Melroy DK. 1982. Polychlorinated Biphenyls in
Human Adipose Tissue and Mother's Milk. Report. Research Triangle
Institute. Research Triangle Park, N.C. RTI/1864/50-03F. NTIS PB83-
253179.
Luotamo M, Jrvisalo, Aitio A. 1985. Analysis of polychlorinated
biphenyls (PCBs) in human serum. Environ Health Perspect 60:327-332.
Lyman VJ, Reehl WF, Rosenblatt DH. 1982. Handbook of Chemical Property
Estimation Methods. New York: McGraw-Hill, pp. 15-16.
-------
120 Section 10
Massachusetts Department of Public Health. 1987. Final Report of Greater
New Bedford PCB Health Effects Study 1984-1987. Boston, Mass: Mass.
Dept. of Public Health.
Mabey WR, Smith JH, Podoll RT, et al. 1981. Aquatic Fate Process Data
for Organic Priority Pollutants. EPA 440/4-81-014. Washington D.C.:
Environmental Protection Agency, Monitoring and Data Support Division,
Office of Water Regulations and Standards, pp. 115-128.
MacLeod KE. 1981. Polychlorinated biphenyls in indoor air. Environ Sci
Technol 15:926-8.
Makiura S, Aoe H. Sugihara S, Hirao K, Arai M, Ito N. 1974. Inhibitory
effect of polychlorinated biphenyls on liver tumorigenesis in rats
treated with 3'-methyl-4-dimethylaminoazobenzene, N-2-
fluorenylacetamide, and diethylnitrosamine. J Nat Cancer Inst 53:1253-
1257 (cited in IARC 1978).
Maroni N, Columbi A, Arbosti G, Cantoni S, Foa V. 1981a. Occupational
exposure to polychlorinated biphenyls in electrical workers. II. Health
effects. Br J Ind Med 38:55-60.
Maroni N, Columbi A, Cantoni S, Ferioli E, Foa V. 1981b. Occupational
exposure to polychlorinated biphenyls in electrical workers. I.
Environmental and blood polychlorinated biphenyls concentrations. Br J
Ind Med 38:49-54.
Masuda Y, Kagawa R, Kuroki H, Tokudom S, Kuratsune M. 1979. Transfer of
various polychlorinated biphenyls to the fetuses and offspring of mice.
Food Cosmet Toxicol 17(6):623-627.
Mazurek MA, Simoneit BRT. 1985. Organic components in bulk and wet-only
precipitation. CRC Crit Rev Environ Control 16:41-47 (cited in EPA
1988a).
McConnell EE, McKinney JD. 1978. Exquisite toxicity in the guinea pig to
structurally-similar halogenated dioxins, furans, biphenyls and
naphthalenes. Toxicol Appl Pharmacol 45:298 (cited in EPA 1988a).
McNulty WP, Becker GM, Cory HT. 1980. Chronic toxicity of 3,3'4,4'- and
2 .2'5,5'-tetrachlorobiphenyls in rhesus macques. Toxicol Appl Pharmacol
56(2):182-190 (cited in EPA 1985a).
Meigs JW, Albom JJ, Kartin Bl. 1954. Chloracne from an unusual exposure
to Arochlor. J Am Med Assoc 154:1417-1418.
Mes J. Doyle JA, Adam BR. Davies DJ, Turton D. 1984. Polychlorinated
biphenyls and organochlorine pesticides in milk and blood of Canadian
women during lactation. Arch Environ Contam Toxicol 13:217-223.
-------
References 121
Mleure JP, Hicks 0, Kaley RG, Saeger VW. 1976. Characterization of
polychlorinated biphenyls. In: National Conference on Polychlorinated
Biphenyls, Chicago, 1975. EPA-560/6-75-004. Washington, D.C.:
Environmental Protection Agency, pp. 84-93.
Miller JV. 1985. Congenital PCB poisoning: A reevaluation. Environ
Health Perspect 60:211-214.
Hizutani T, Hidaka K, Matsumoto H. 1977. A comparative study on
accumulation and elimination of tetrachlorobiphenyl isomers in mice.
Bull Environ Contain Toxicol 18:454 (cited in EPA 1988a) .
Monsanto. 1974. PCBs-Aroclors Tech Bull. 0/PL 306A. St. Louis, Mo (cited
in Callahan et al. 1979)
Morgan RW, Ward JM, Hartman PE. 1981. Aroclor 1254-induced intestinal
metaplasia and adenocarcinoma in the glandular stomach of F344 rats.
Cancer Res 41:5052-5059.
Morselli L, Brocco D, Pirni A. 1985. The presence of polychlorodibenzo-
p-dioxins (PCDDs), polychlorodibenzofurans (PCDFs), and
polychlorobiphenyls (PCBs) in fly ashes from various municipal
incinerators under different technological and working conditions. Ann
Chim 75:59-64.
Mosley CL, Emmett E. 1984. NIOSH Health Hazard Evaluation Report. GSA.
Switchgear Shop. HETA80-007-1520. NTIS PB86-133741.
Muehlebach S, Bickel MH. 1981. Pharmacokinetics in rats of
2,4,5,2',4',5'-hexachlorobiphenyl, an unmetabolizable lipophilic model
compound. Xenobiotica 11(4):249-257 (cited in EPA 1988a).
Murphy TJ, Formanski LJ. Brovnawell B, Meyer JA. 1985. Polychlorinated
biphenyl emissions to the atmosphere in the Great Lakes region.
Municipal land fills and incinerators. Environ Sci Technol 19(10):924-
946.
Murray HE, Ray LE, Giaa CS. 1981. Phthalic acid esters, total DDT and
polychlorinated biphenyls in marine samples from Calveston Bay, Texas.
Bull Environ Contam Toxicol 26:769-774.
Nagasaki H, Tomii S, Mega T. 1975. Factors affecting induction of liver
cancer by BHC and PCBs in mice. Abstract No. 235. Jpn J Hyg 30:134
(cited in IARC 1978).
NAS (National Academy of Sciences). 1980. Drinking Water and Health.
Vol. 3, Washington, D.C.: National Academy Press, pp. 25-67 (cited in
EPA 1988a).
NCI (National Cancer Institute). 1978. Bioassay of Aroclor 1254 for
possible carcinogenic!ty. NCI-GC-TR-38. Betheseda, Md: National Cancer
Institute. NTIS PB279624.
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122 Section 10
Needham LL, Amrek AL, Head SL, Burse Vtf and Llddle JA. 1980. Column
chromatography separation of polychlorinated biphenyls from
dlchlorodiphenyltrichloroethane and metabolites. Anal Chem 52:2227-2229.
Needhaa LL, Burse Vtf, Price HA. 1981. Temperature-programmed gas
chromatographie determination of polychlorinated and polybrominated
biphenyls in serum. J Assoc Off Anal Chem 64:1131-1137.
Nelson NN, Hammon PB, Nisbet ICT, Sarofim AF. Drury UH. 1972.
Polychlorinated biphenyls - environmental impact. Environ Res 5:249-362
(cited in EPA 1988a).
Nilsson B, Ramel C. 1974. Genetic tests on Drosophila melanogaster with
polychlorinated biphenyls (PCB). Hereditas 77:319-322 (cited in EPA
1988a).
NIOSH (National Institute for Occupational Safety and Health). 1977a.
NIOSH Manual of Analytical Methods. 2nd ed. Taylor DG, ed. Vol. 1.
Cincinnati, Ohio: U.S. Department of Health and Human Services, NIOSH,
pp. 244-1 - 253-7.
NIOSH (National Institute for Occupational Safety and Health). 1977b.
Criteria for a recommended standard. Occupational Exposure to
Polychlorinated Biphenyls (PCBs). Rockville, Md: U.S. Department of
Health, Education and Welfare, Public Health Service, Centers for
Disease Control. NIOSH Publ 77-225.
NIOSH (National Institute for Occupational Safety and Health). 1984a.
NIOSH Manual of Analytical Methods. 3rd ed. Eller PM. ed. Vol. 2.
Cincinnati, Ohio: U.S. Department of Health and Human Services, NIOSH,
pp. 5503-1 - 5503-5.
NIOSH (National Institute for Occupational Safety and Health). 1984b.
NIOSH Manual of Analytical Methods. 3rd ed. Eller PM, ed. Vol. 1.
Cincinnati, Ohio: U.S. Department of Health and Human Services, NIOSH,
pp. 8004-1 - 8004-4.
NIOSH (National Insitute for Occupational Safety and Health). 1986.
Polychlorinated Biphenyls (PCBS): Potential Health Hazards from
Electrical Equipment Fires or Failures. Department of Health and Human
Services. NIOSH Publ 86-111.
Nishizumi M. 1976. Radioautographic evidence for adsorption of
polychlorinated biphenyls through the skin. Ind Health 14:41-44.
Norback DH, Mack E. Blomquise KA, Allen JR. 1978. Metabolic study of
2,4,5,2',4',5'-hexachlorobiphenyl in rhesus monkeys. Toxicol Appl
Pharmacol 45:331 (cited in EPA 1985a).
* Norback DH. Weltnan RH. 1985. Polychlorinated biphenyl Induction of
hepatocellular carcinoma in the Sprague-Dawley rat. Environ Health
Perspect 60:97-105.
-------
References 123
NTIS (National Technical Information Service). 1987. Federal research in
progress: On-line database.
Nyhan WL. 1961. Toxicity of drugs in the neonatal period. J Pediatr
59(1):1 (cited in EPA 1988a).
Oatman L, Roy R. 1986. Surface and indoor air levels of polychlorinated
biphenyls in public buildings. Bull Environ Contain Toxicol 37:461-466.
Oesterle D, Demi E. 1983. Promoting effect of polychlorinated biphenyls
on development of enzyme-altered islands in livers of weanling and adult
rats. J Cancer Res Clin Oncol 105(2):141-146 (cited in EPA 1985a).
Orris P, Kominsky JR, Hryhorczyk D, Melius J. 1986. Exposure to
polychlorinated biphenyls from an overheated transformer. Chemosphere
15:1305-1311.
OSHA (Occupational Safety and Health Administration). 1985. Code of
Federal Regulations. OSHA Occupational Standards. Permissible Limits.
29 CFR 1910.1000.
Ouw HK, Simpson GR, Siyali DS. 1976. Use and health effects of Aroclor
1242, a polychlorinated biphenyl in an electrical industry. Arch Environ
Health 31:189.
Paris DF, Steen VC, Baughman GL. 1978. Role of the physicochemical
properties of Aroclor 1016 and 1242 in determining their fate and
transport in aquatic environments. Chemosphere 7(4):319-325 (cited in
Callahan et al. 1979)
Parkinson A, Thomas PE, Ryan DE, et al. 1983. Differential time course
of induction of rat liver microsomal cytochrom P-450 isozymes and
epoxide hydrolase by Aroclor 1254. Arch Biochem Biophys 225:203-215
(cited in EPA 1988a).
Peakall DB, Lincer JL, Bloom SE. 1972. Embryonic mortality and
chromosomal alterations caused by Aroclor 1254 in ring doves. Environ
Health Perspect 1:103-104 (cited in EPA 1988a).
Pereira MA, Herren SL, Britt AL, Khoury MM. 1982. Promotion by
polychlorinated biphenyls of enzyme-altered foci in rat liver. Cancer
Lett 15(2):185-190 (cited in EPA 1985a).
Ray LE, Murray HE, Giam CS. 1983. Organic pollutants in marine samples
from Portland, Maine. Chemosphere 12:1031-1038.
Reid D, Fox JM. 1982. Polychlorinated biphenyl report, Old Forge.
Lackavanna County, Pennsylvania Department of Health, Division of
Environmental Health, April 1982 (cited in Kreiss 1985).
Requejo AG, Vest RH, Hatcher PG. McGillivary PA. 1979. Polychlorinated
biphenyls and chlorinated pesticides in soils of the Everglades national
park and adjacent agricultural areas. Environ Sci Technol 13:931-936.
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124 Section 10
* Ringer RK, Aulerich RJ, Bleavlns MR. 1981. Biological effects of PCBs
and PBBs on mink and ferrets: A review. In: Khan MAQ, ed. Halogenated
Hydrocarbons: Health and Ecological Effects. Elmsford N.Y.: Permagon
Press, pp. 329-343 (cited in Homshaw et al. 1986).
Rodgers PU, Swain UR. 1983. Analysis of polychlorinated biphenyl (PCB)
loading trends in Lake Michigan. J Great Lakes Res 9:548-58.
Rogan WJ, Gladen BC, McKinney JD, et al. 1986. Neonatal effects of
transplacental exposure to PCBs and DDE. J Pediatr 109:335-341.
Rogan WJ, Gladen BC, McKinney JD et al. 1987. Polychlorinated biphenyls
(PCBs) and dichlorodiphenyl dichloroethane (DDE) in human milk: Effects
on growth, morbidity and duration of lactation. Am J Public Health
77:1294-1297.
Rozanova LF. 1943. [Toxicity of some chlorinated aromatic hydrocarbons].
Farmakol Toksikol 6:48. (In Russian) (cited in NIOSH 1977b).
Ryan JJ, Lau PY, Pilon JC, Lewis D, McLeod HA, Gervais A. 1984.
Incidence and levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin in Lake
Ontario commercial fish. Environ Sci Techno1 18:719-721.
Safe S. 1976. Overview of analytical identification and spectroscopic
properties. In: National Conference on Polychlorinated Biphenyls,
Chicago, 1975. EPA-560/6-75-004. Washington, D.C.: Environmental
Protection Agency, pp. 94-102 (cited in IARC 1978).
Safe S. 1980. Affadivit dated April 23, 1980. University of Guelph.
Guelph, Ontario, Canada.
Safe S. 1980. Halogenated biphenyls, terphenyls, napthalenes,
dibenzodioxins and related products. Metabolism uptake, storage and
bioaccumulation. Toxicol Environ Health 4:81-107.
Safe S, Bandiera S, Sawyer T, et al. 1985a. PCBs: Structure-function
relationships and mechanism of action. Environ Health Perspect 60:47-56.
Safe S, Hutzinger 0, Jones D. 1975. The mechanism of chlorobiphenyl
metabolism. J Agric Food Chem 23:851-853.
Safe S, Safe L, Mullin M. 1985b. Polychlorinated biphenyls: Congener-
specific analysis of a commercial mixture and a human milk extract. J
Agric Food Chem 33:24-29.
Sager DB. 1983. Effect of postnatal exposure to polychlorinated
biphenyls on adult male reproductive function. Environ Res 31(1):76-94.
Sahl JD, Crocker IT, Gordon RJ, Faeder EJ. 1985a. Polychlorinated
biphenyl concentrations in the blood plasma of a selected sample of
nonoccupationally exposed Southern California working adults. Sci Total
Environ 46:9-18.
-------
References 125
Sahl JD, Crocker T, Gordon RJ, Faeder EJ. 1985b. Polychlorinated
biphenyls in Che blood of personnel from an electric utility. J Occup
Hed 27:639-643.
* Sanders OT, Zepp RL, Kirkpatrick RL. 1974. Effect of PCB ingestion on
sleeping times, organ weights, food consumption, serum corticosterone
and survival of albino mice. Bull Environ Contain Toxicol 12(4):394-399.
SANSS (Structure and Nomenclature Search System). 1987. Chemical
Information System (CIS) computer data base.
Schaeffer E, Greim H, Goessner W. 1984. Pathology of chronic
polychlorinated biphenyl (PCB) feeding in rats. Toxicol Appl Pharmacol
75:278-288.
Schecter A, Tiernan T. 1985. Occupational exposure to polychlorinated
dioxins, polychlorinated furans, polychlorinated biphenyls, and
biphenylenes after an electrical panel and transformer accident in an
office building in Binghamton, N.Y. Environ Health Perspect 60:305-313.
Schmitt CJ, Zajicek JL, Ribick MA. 1985. National pesticide monitoring
program. Residues of organochlorine chemicals in freshwater fish, 1980-
1981. Arch Environ Contam Toxicol 14:225-260.
Schneider JF, Bourne S, Boparai S. 1984. Parallel capillary column gas
chromatography in the determination of chlorinated pesticides and PCBs.
J Chromatogr 22(5):203-206.
Schnellmann RG, Putnam CV, Sipes IG. 1983. Metabolism of
2.2',3,3',6,6'-hexachlorobiphenyl and 2,2',4,4',5,5'-hexachlorobiphenyl
by human hepatic microsomes. Biochem Pharmacol 32:3233-3239 (cited in
EPA 1988a).
Schnellmann RG, Volp RF, Putnam CW, Sipes IG. 1984. The hydroxylation,
dechlorination and glucuronidation of 4,4'-dichlorobiphenyl by human
hepatic microsomes. Biochem Pharmacol 33:3503-3509 (cited in EPA 1988a).
Schoeny R. 1982. Mutagenicity testing of chlorinated biphenyls and
chlorinated dibenzofurans. Mutat Res 101:45-56 (cited in EPA 1988a).
Schoeny RS, Smith CC, Loper JC. 1979. Non-mutagenicity for Salmonella of
the chlorinated hydrocarbons Aroclor 1254, 1,2,4-trichlorobenzene, mirex
and kepone. Mutat Res 68:125.
Schwartz PM, Jacobson SV, Fein G, Jacobson JL. Price HA. 1983. Lake
Michigan fish consumption as a source of polychlorinated biphenyls in
human cord serum, maternal serum and milk. Pub Amer J Public Health
73(3):293-296.
Sipes IG, McLain GE, Jr, Podolsky TL, Brown BR, Jr. 1978. Bioactivation
of halothane: Correlation with hepatotoxicity. Int Congr Serx-Excerpta
Med 440:238 (cited in EPA 1985a).
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126 Section 10
Sklarew DS, Glrvln DC. 1987. Attenuation of polychlorinated biphenvls ir
soils. Rev Environ Contain Toxicol 98:1-41.
Smith AB, Schloemer J, Lowry LK, et al. 1981a. Cross-Sectional Medical
Survey of a Group of Workers Occupationally Exposed to Polychlorinated
Biphenyls (PCBs) at an Electrical Equipment Manufacturing Plant.
Cincinnati, Ohio: National Institute for Occupational Safety and Health
Division of Surveillance, Hazard Evaluations and Field Studies and
Lipid Research Center, University of Cincinnati Medical Center'(cited in
Drill et al. 1981).
Smith AB, Schloemer J, Lowry LK, et al. 1981b. Cross-Sectional Medical
Survey of Two Groups of Workers Occupationally Exposed to
Polychlorinated Biphenyls (PCBs) in the Maintenance, Repair, and
Overhaul of Electrical Transformers. Cincinnati, Ohio: National
Institute for Occupational Safety and Health, Division of Surveillance
Hazard Evaluations and Field Studies, and Lipid Research Center,
University of Cincinnati Medical Center (cited in Drill et al. 1981).
Smith AB, Schloemer J, Lowry LK, et al. 1981c. Metabolic and Health
Consequences of Occupational Exposure to Polychlorinated Biphenyls
(PCBs). Cincinnati, Ohio: National Institute for Occupational Safety and
Health, Division of Surveillance, Hazard Evaluations and Field Studies,
and Lipid Research Center, University of Cincinnati Medical Center
(cited in Drill et al. 1981).
Smith AB, Schloemer J, Lowry LK, et al. 1982. Metabolic and health
consequences of occupational exposure to polychlorinated biphenyls. Br J
Ind Med 39:361-369 (cited in Wolff 1985, Kreiss 1985).
Smrek AL, Needham LL. 1982. Simplified cleanup procedures for adipose
tissue containing polychlorinated biphenyls, DDT, and DDT metabolites.
Bull Environ Contain Toxicol 28:718-722.
Sparling J, Fung D, Safe S. 1980. Bromo- and chlorobiphenyl metabolism:
GC/MS identification of urinary metabolites and the effects of structure
on their rates of excretion. Biomed Mass Spectrom 7:13-20 (cited in EPA
1988a).
Spencer F. 1982. An assessment of the reproductive toxic potential of
Aroclor 1254 in female Sprague-Dawley rats. Bull Environ Contain Toxicol
28(3):290-297.
Steinberg KK, Frenl-Titulaer LWJ, Rogers TO, et al. 1986. Effects of
polychlorinated biphenyls and lipemia on serum analytes. J Toxicol
Environ Health 19:369-381.
Stone PJ Ed. 1981. Emergency Handling of Hazardous Materials in Surface
Transportation. Washington, D.C.: Bureau of Explosives, Association of
American Railroads, p. 418.
Sundstrom G, Hutzinger D, Safe S. 1976a. The metabolism of
chlorobephenyls - A review. Chemosphere 5:267.
-------
References 127
Sundstrom G, Hutzinger D, Safe S. 1976b. The metabolism of
2,2'.4,4',5,5'-hexachlorobiphenyl by rabbits, rats and mice. Chemosphere
4:249 (cited in EPA 1988a).
Swackhamer DL, Armstrong DE. 1986. Estimation of the atmospheric and
nonatmospheric contributions and losses of polychlorinated biphenyls for
Lake Michigan on the basis of sediment records of remote lakes. Environ
Sci Technol 20(9):879-883.
Tanabe S, Nakagawa Y, Tatsukawa R. 1981. Absorption efficiency and
biological half-life of individual chlorobiphenyls in rats treated
chlorobiphenyl products. Agric Biol Chem 45:717-726 (cited in EPA
1988a).
Tanabe S, Hidaka H, Tatsukawa R. 1983. PCBs and chlorinated hydrocarbon
pesticides in Antarctic atmosphere and hydrosphere. Chemosphere 12:277-
288.
Tanabe S, Tanaka H. Tatsukawa R. 1984. Polychlorobiphenyls, DDTs and
hexachlorocyclohexane isomers in the western North Pacific ecosystem.
Arch Environ Contam Toxicol 13:731-738.
Tatematsu M, Nakanishi K, Murasaki G, Miyata Y, Hirose M, Ito N. 1979.
Enhancing effect of inducers of liver microsomal enzymes on induction of
hyperplastic liver nodules by tf-2-fluorenylacetamide in rats. J Nat
Cancer Inst 63(6):1411-1416 (cited in EPA 1985a).
Taylor PR. Lawrence CE. Hwang HL, Paulson AS. 1984. Polychlorinated
biphenyls: Influence on birthweight and gestation. Am J Public Health
74(10):1153-1154.
* Thomas PT, Hinsdill RD. 1978. Effect of polychlorinated biphenyls on
the immune responses of rhesus monkeys and mice. Toxicol Appl Pharmacol
44:41-51.
Thomas RL, Frank R. 1981. PCBs in sediment and fluvial suspended solids
in the Great Lakes. In: Mackay D, et al., eds. Phys Behav PCDs Great
Lakes. Ann Arbor, Mich: Ann Arbor Science, pp. 245-267.
Tiernan TO, Taylor ML, Garret JH, et al. 1983. PCDDs, PCDFs and related
compounds in the effluents from combustion processes. Chemosphere
12:595-606.
Tiernan TO, Taylor ML, Garret JH, et al. 1985. Sources and fate of
polychlorinated dibenzodioxina, dibenzofurans and related compounds in
human environments. Environ Health Perspect 59:145-58.
* Treon JF, Cleveland FP, Cappel JW, Atchley RW. 1956. The toxicity of
the vapours of Aroclor 1242 and Aroclor 1254. Am Ind Hyg Assoc Q
17:204-213.
-------
128 Section 10
USITC (U.S. International Trade Commission). 1978. Imports of benzenoid
chemicals and products 1977. USITC Publ 900. Washington, D.C.: USITC
p. 26. i
USITC (U.S. International Trade Commission). 1979. Imports of benzenoid
chemicals and products 1978. USITC Publ 990. Washington, D.C.: USITC,
p. 26.
USITC (U.S. International Trade Commission). 1980. Imports of benzenoid
chemicals and products 1979. USITC Publ 1083. Washington, D.C.: USITC,
p. 28.
USITC (U.S. International Trade Commission). 1982. Imports of benzenoid
chemicals and products 1981. USITC Publ 1272. Washington, D.C.: USITC,
p. 25.
Van Duuran BL. 1981. Cocarcinogens and tumor promoters and their
environmental importance. J Environ Pathol Toxicol 4:959-960.
Vernon AA, et al.** 1981. High levels of polychlorinated biphenyls in
serum specimens, Kansas. Internal report ELI-80-23-2, Centers for
Disease Control, Atlanta, November 16, 1981 (cited in Kreiss 1985).
View data base. 1989. Agency for Toxic Substances and Disease Registry
(ATSDR). Atlanta, Georgia: Office of External Affairs, Exposure and
Disease Registry Branch, February 1989.
* Villeneuve DC, Grant DL, Khera K, Clegg DJ, Baer H, Phillips WEJ.
1971. The fetotoxicity of a polychlorinated biphenyl mixture (Aroclor
1254) in the rabbit and in the rat. Environ Physiol 1:67-71.
Villeneuve DC, Grant DL, Phillips WEJ. 1972. Modification of
pentobarbital sleeping times in rats following chronic PCB ingestion
Bull Environ Contam Toxicol 7:264 (cited in EPA 1985a).
* Vos JG, Beems RB. 1971. Dermal toxicity studies of technical
polychlorinated biphenyls and fractions thereof in rabbits. Toxicol Apol
Pharmacol 19:317-633.
Vos JG, Notenboom-Ram E. 1972. Comparative toxicity study of 2,4,5,
2',4',5'-hexachlorobiphenyl and a polychlorinated biphenyl mixture in
rabbits. Toxicol Appl Pharmacol 23:563-578.
Vos JG, deRoiJ T. 1972. Immunosuppressive activity of a polychlorinated
biphenyl preparation on the humoral immune response in guinea pies
Toxicol Appl Pharmacol 21:549-555.
Vos JG, van Genderen H. 1973. Toxicological aspects of
immunesuppressIon. In: Deichaan WB, ed. Pesticides in the Environment A
Continuing Controversy. Miami, Fla.: Eighth International Conference on
Toxicology and Occupational Medicine. New York: Intercontinental Medical
Book Co. (cited in EPA 1988a).
-------
References 129
Ward JM. 1985. Proliferative lesions of the glandular stomach and liver
In F344 rats fed diets containing Arclor 1254. Environ Health Perspect
60:89-95.
Watanabe M, Sugahara T. 1981. Experimental formation of cleft palate in
mice with polychlorinated biphenyls (PCB). Toxicology 19(l):49-53 (cited
in EPA 1988a).
Weant GE, McCormick GS. 1984. Nonindustrial sources of potential toxic
substances and their applicability to source apportionment methods. EPA
450/4-84-003; NTIS PB84-231232. Research Triangle Park, N.C.:
Environmental Protection Agency, pp. 36, 86.
Welsch F. 1985. Effects of acute or chronic polychlorinated biphenyl
ingestion on maternal metabolic homeostasis and on the manifestations of
embryotoxicity caused by cyclophosphamide in mice. Arch Toxicol
27(2):104-113.
Welty ER. 1983. Personal communication, August 8, 1983 (cited in Kreiss
1985).
Wester RC, Bucks DAW, Haibach HI, Anderson J. 1983. Polychlorinated
biphenyls (PCBs): Dermal absorption, systemic elimination and dermal
wash efficiency. J Toxicol Environ Health 12:511-519.
Vickizer TM, Brilliant LB, Copeland R, Tilden R. 1981. Polychlorinated
biphenyl contamination of nursing mothers' milk in Michigan. Am J Public
Health 71:132-137.
Wolff MS. 1983. Occupational derived chemicals in breast milk. Am J Ind
Med 4:259-281 (cited in EPA 1985a).
Wolff MS. 1985. Occupational exposure to polychlorinated biphenyls
(PCBs). Environ Health Perspect 60:133-138.
Wolff MS, Fischbein A, Thornton J, Rice C, Lillis R, Selikoff IJ. 1982a.
Body burden of polychlorinated biphenyls among persons employed in
capacitor manufacturing. Int Arch Occup Environ Health 49:199-208 (cited
in Kreiss 1985).
Wolff MS, Thornton J, Fischbein A, Lillis R, Selikoff IJ. 1982b.
Disposition of polychlorinated biphenyl congeners in occupationally
exposed person. Toxicol Appl Fharmacol 62(2):294-306.
Wyndham C, Devenish J, Safe S. 1976. The in vitro metabolism,
macromolecular binding and bacterial mutagenicity of 4-chlorobiphenyl, a
model PCB substrate. Res Commun Chem Pathol Pharmacol 15:563 (cited in
EPA 1985a)
Wyss PA, Muhleback S, Bickel MH. 1986. Long-term pharmacokinetics of
2,2',4,4*,5,5'-hexachlorobiphenyl (6-CB) in rats with constant adipose
tissue mass. Drug Metab Dispos 14:361-365 (cited in EPA 1988a).
-------
130 Section 10
Yakushljl T, Watanabe I, Kuwabara et al. 1978. Long-term studies of the
excretion of poly chlorinated bipheyls (PCBs) through the mother's milk
of an occupationally-exposed worker. Arch Environ Toxicol 7:493-504
(cited in EPA 1988a).
Yoshimura H, Yamamoto HA. 1975. A novel route of excretion of
2,4,3',4'-tetrachlorobiphenyl in rats. Bull Environ Contain Toxicol
13:681-688 (Cited in EPA 1985a).
Yoshimura H, Yoshihara S. 1976. Toxicological aspects. II. The metabolic
fate of PCBs and their toxicological evaluation. In: Higuchi K. ed. PCB
Poisoning and Pollution. Tokyo: Kondansha Ltd, pp. 41-67 (cited in EPA
1985a)
Young SS. 1985. Letter to the editor. Toxicol Appl Pharmacol 78:321-322.
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131
11. GLOSSARY
Acute Exposure--Exposure to a chemical for a duration of 14 days or
less, as specified In the Toxlcologlcal Profiles.
Bloconcentratlon Factor (BCP)--The quotient of the concentration of a
chemical In aquatic organisms at a specific time or during a discrete
time period of exposure divided by the concentration in the surrounding
water at the same time or during the sane time period.
Carcinogen--A chemical capable of inducing cancer.
Ceiling value (CL)--A concentration of a substance that should not be
exceeded, even instantaneously.
Chronic Exposure--Exposure to a chemical for 365 days or more, as
specified in the Toxicological Profiles.
Developmental Toxicity--The occurrence of adverse effects on the
developing organism that may result from exposure to a chemical prior to
conception (either parent), during prenatal development, or postnatally
to the time of sexual maturation. Adverse developmental effects may be
detected at any point in the life span of the organism.
Embryotoxicity and Fetotoxicity--Any toxic effect on the conceptus as a
result of prenatal exposure to a chemical; the distinguishing feature
between the two terms is the stage of development during which the
insult occurred. The terms, as used here, include malformations and
variations, altered growth, and in utero death.
Frank Effect Level (FEL)--That level of exposure which produces a
statistically or biologically significant increase in frequency or
severity of unmistakable adverse effects, such as irreversible
functional Impairment or mortality, in an exposed population when
compared with its appropriate control.
EPA Health Advisory--An estimate of acceptable drinking water levels for
a chemical substance baaed on health effects information. A health
advisory is not a legally enforceable federal standard, but serves as
technical guidance to assist federal, state, and local officials.
idiately Dangerous to Life or Health (XDLH)-*The maximum
environmental concentration of a contaminant from which one could escape
within 30 min without any escape-Impair ing symptom* or irreversible
health effects.
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132 Section 11
Intermediate Exposure--Exposure to a chemical for a duration of 15-364
days, as specified in the Toxicological Profiles.
Immunologic Toxicity--The occurrence of adverse effects on the immune
system that may result from exposure to environmental agents such as
chemicals.
In vitro--Isolated from the living organism and artificially maintained,
as in a test tube.
In vivo--Occurring within the living organism.
Key Study--An animal or human toxicological study that best illustrates
the nature of the adverse effects produced and the doses associated with
those effects.
Lethal Coneentration(LO) (LCLO)--The lowest concentration of a chemical
in air which has been reported to have caused death in humans or
animals.
Lethal Concentration(SO) (LCso)--A calculated concentration of a
chemical in air to which exposure for a specific length of time is
expected to cause death in 50% of a defined experimental animal
population.
Lethal Dose(LO) (LDLO)--The lowest dose of a chemical introduced by a
route other than inhalation that is expected to have caused death in
humans or animals.
Lethal Dose(50) (LDso)--The dose of a chemical which has been calculated
to cause death in 50% of a defined experimental animal population.
Lowest-Observed-Adverse-Effect Level (LOAEL)--The lowest dose of
chemical in a study or group of studies which produces statistically or
biologically significant increases in frequency or severity of adverse
effects between the exposed population and its appropriate control.
Lowest-Observed-Effect Level (LOEL)--The lowest dose of chemical in a
study or group of studies which produces statistically or biologically
significant increases in frequency or severity of effects between the
exposed population and its appropriate control.
Malformations--Permanent structural changes that may adversely affect
survival, development, or function.
Minimal Risk Level--An estimate of daily human exposure to a chemical
that is likely to be without an appreciable risk of deleterious effects
(noncancerous) over a specified duration of exposure.
Mutagen--A substance that causes mutations. A mutation is a change in
the genetic material in a body cell. Mutations can lead to birth
defects, miscarriages, or cancer.
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Glossary 133
Neurotoxiclty- -The occurrence of adverse effects on the nervous system
following exposure to a chemical.
No-Observed-Adverse-Effect Level (NOAEL)--That dose of chemical at which
there are no statistically or biologically significant increases in
frequency or severity of adverse effects seen between the exposed
population and its appropriate control. Effects may be produced at this
dose, but they are not considered to be adverse.
No-Observed-Effect Level (NOEL) --That dose of chemical at which there
are no statistically or biologically significant increases in frequency
or severity of effects seen between the exposed population and its
appropriate control.
Permissible Exposure Limit (PEL) --An allowable exposure level in
workplace air averaged over an 8-h shift.
q^-'The upper-bound estimate of the low-dose slope of the dose-response
curve as determined by the multistage procedure. The q * can be used to
calculate an estimate of carcinogenic potency, the incremental excess
cancer risk per unit of exposure (usually /ig/L for water, mg/kg/day for
food, and /*g/m3 for air).
Reference Dose (RfD)--An estimate (with uncertainty spanning perhaps an
order of magnitude) of the daily exposure of the human population to a
potential hazard that is likely to be without risk of deleterious
effects during a lifetime. The RfD is operationally derived from the
NOAEL (from animal and human studies) by a consistent application of
uncertainty factors that reflect various types of data used to estimate
RfDs and an additional modifying factor, which is based on a
professional judgment of the entire database on the chemical. The RfDs
are not applicable to non threshold effects such as cancer.
Reportable Quantity (RQ)--The quantity of a hazardous substance that is
considered reportable under CERCLA. Reportable quantities are: (1) 1 Ib
or greater or (2) for selected substances, an amount established by
regulation either under CERCLA or under Sect. 311 of the Clean Water
Act. Quantities are measured over a 24-h period.
Reproductive Toxiclty--The occurrence of adverse effects on the
reproductive system that may result from exposure to a chemical . The
toxic ity may be directed to the reproductive organs and/or the related
endocrine system. The manifestation of such toxicity may be noted as
alterations in sexual behavior, fertility, pregnancy outcomes, or
modifications in other functions that are dependent on the integrity of
this system.
Short -Ten Exposure Limit (STEL)--The maximum concentration to which
workers can be exposed for up to 15 min continually. No more than four
excursions are allowed per day, and there must be at least 60 min
between exposure periods. The daily TLV-TWA may not be exceeded.
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134 Section 11
Target Organ Toxlclty--This term covers a broad range of adverse effects
on target organs or physiological systems (e.g., renal, cardiovascular)
extending from those arising through a single limited exposure to those
assumed over a lifetime of exposure to a chemical.
Teratogen--A chemical that causes structural defects that affect the
development of an organism.
Threshold Limit Value (TLV)--A concentration of a substance to which
most workers can be exposed without adverse effect. The TLV may be
expressed as a TWA. as a STEL, or as a CL.
Time-weighted Average (TWA)--An allowable exposure concentration
averaged over a normal 8-h workday or 40-h workweek.
Uncertainty Factor (UF)--A factor used in operationally deriving the RfD
from experimental data. UFs are intended to account for (1) the
variation in sensitivity among the members of the human population,
(2) the uncertainty in extrapolating animal data to the case of humans,
(3) the uncertainty in extrapolating from data obtained in a study that
is of less than lifetime exposure, and (4) the uncertainty in using
LOAEL data rather than NOAEL data. Usually each of these factors is set
equal to 10.
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135
APPENDIX: PEER REVIEW
A peer review panel was assembled for PCBs. The panel consisted of
the following members: Dr. Rolf Hartung, Chairman, Toxicology Program,
University of Michigan; Dr. James Olson, Associate Professor of
Pharmacology and Therapeutics, SUNY Buffalo; Dr. Shane Que Hee,
Associate Professor of Environmental Health, University of Cincinnati
Medical Center. These experts collectively have knowledge of PCB's
physical and chemical properties, toxicokinetics, key health end points,
mechanisms of action, human and animal exposure, and quantification of
risk to humans. All reviewers were selected in conformity with the
conditions for peer review specified in the Superfund Amendments and
Reauthorization Act of 1986, Section 110.
A Joint panel of scientists from ATSDR and EPA has reviewed the
peer reviewers' comments and determined which comments will be included
in the profile. A listing of the peer reviewers' comments not
incorporated in the profile, with a brief explanation of the rationale
for their exclusion, exists as part of the administrative record for
this compound. A list of databases reviewed and a list of unpublished
documents cited are also included in the administrative record.
The citation of the peer review panel should not be understood to
imply their approval of the profile's final content. The responsibility
for the content of this profile lies with the Agency for Toxic
Substances and Disease Registry.
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