POLYCHLORINATED BIPHENYLS
Ambient Water Quality Criteria
Criteria and Standards Division
Office of Water Planning and Standards
U.S. Environmental Protection Agency
Washington, D.C.
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CRITERION DOCUMENT
POLYCHLORINATED BIPHENYLS
Criteria
Aquatic Life
For polychlorinated biphenyls the criterion to protect
i
freshwater aquatic life as derived using the Guidelines
is 0.0015 jug/1 as a 24-hour average and the concentration
should not exceed 6.2 ;ug/l at any time.
For polychlorinated biphenyls the criterion to protect
saltwater aquatic life as derived using the Guidelines is
0.024/ig/l as a 24-hour average and the concentration should
not exceed 0.20 ;ag/l at any time.
Human Health
For the maximum protection of human health from the
potential carcinogenic effects of exposure to PCBs through
ingestion of water and contaminated aquatic organisms, the
ambient water concentration should be zero. Concentrations
of PCBs estimated to result in additional lifetime cancer
risks ranging from no additional risk to an additional risk
of 1 in 100,000 are presented in the Criterion Formulation
section of this document. The Agency is considering setting
criteria at an interim target risk level in the range of
10" , 10 , or 10~ with corresponding criteria of 0.26 ng/1,
0.026 ng/1 and 0.0026 ng/1 respectively.
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Introduction
Polychlorinated biphenyls (PCBs) are the chlorinated
derivatives of a class of aromatic organic compounds called
biphenyls and are manufactured by the direct chlorination
of the biphenyl ring system. The commercial products are
complex mixtures of chlorobiphenyls and are marketed for
various uses according to the percentage of chlorine in
the mixture. Currently there is no production of PCBs in
the United States but the sole producer of PCBs in the United
States previously marketed four mixtures containing 21 percent,
41 percent, 42 percent and 54 percent chlorine for use only
in closed electrical systems under the trademark "Aroclor."
Prior to 1971 mixtures containing up to 68 percent chlorine
were used in a number of other applications, including plasti-
cizers, heat transfer fluids, hydraulic fluids, fluids in
vacuum pumps and compressors, lubricants, and wax extenders.
In 1974 approximately 65 to 70 percent of domestic
sales were to manufacturers of capacitors and the remainder
to manufacturers of transformers while approximately 450,000
pounds of PCBs were imported primarily for use in non-closed
systems. U.S. production appeared to be one-half of the
world total.
As a result of the long life of many products containing
PCBs, it is believed that a substantial portion of the PCBs
manufactured before 1971 are still in service and thus represent
potential pollution through possible future discharge into
the environment.
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During the period 1972 to 1974 domestic production of
PCBs averaged approximately 40 million pounds per year with
33 million pounds representing the annual domestic marketed
consumption during that period.
Although the environmental behavior and biological
activity of a number of individual chlorobiphenyl isomers
have been studied in recent years, it is still difficult
to evaluate the potential toxicity of the complex mixtures
actually found in the environment since their composition
often changes. In making this evaluation it is necessary
to weigh carefully the results of studies of individual
compounds, and to compare critically the environmental and
toxicological properties of the commercial mixtures.
A further complication is that several commercial PCB
mixtures have been reported to contain small quantities
of highly toxic contaminants, polychlorinated dibenzofurans
(PCDFs). Certain of the toxic effects observed in animals
and humans exposed to PCBs appear to be attributable to
PCDFs, while others appear to be caused by PCBs themselves.
There is also some evidence that small quantities of PCDFs
may be formed from PCBs while in service or as a result
of metabolic changes in certain organisms.
PCBs consist of a mixture of chlorinated biphenyls
which contain a varying number of substituted chlorine atoms
on the aromatic rings. The biphenyl molecule has a total
of ten sites where chlorine substitution can be accommodated
as shown in the following structure:
A-2
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The potential positions for chlorine substitution are numbered
according to the American Chemical Society standard notation.
Chlorinated biphenyls having the same number of chlorine
• f, • . . •
atoms per molecule are referred to as a specific class of
chlorobiphenyls, with a suitable numerical prefix to define
the number of substituted chlorines. Hence/ there are classes
varying from monochlorobiphenyls to decachlorobiphenyl.
All compounds within the same class have the same molecular
weight and are structural isomers of each other. They differ
only in terms of the location of the chlorine atoms in the
biphenyls ring. The ten classes of chlorobiphenyls, comprissing
209 possible isomers, are summarized in Table 1.
Chlorobiphenyls with five or more chlorine atoms are referred
to as "higher chlorobiphenyls." This distinction is made
in recognition of the fact that the former group of compounds
is much more persistent in the environment than the latter
group. The tetrachlorobiphenyls are intermediate in persistence.
The physical properties of individual chlorinated biphenyls
are known (Cook, 1972). The physical properties of the
Aroclor mixtures are summarized in Table 2. Lower chlorinated
Aroclors (1221, 1232, 1016, 1242, and 1248) are colorless mobile
oils. Increasing chlorine content results in mixtures taking
on the consistency of viscous liquids (Aroclor 1254) or
sticky resins (Aroclors 1260 and 1262). Aroclors 1268 and
1270 are off white-white powders. With the exception of
Aroclors 1221 and 1268, Aroclors do not crystallize upon
- /
A-3
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TABLE 1
Emperical Formulation, Molecular Weights
and Chlorine Percentage in PCBs
Empirical formula
chlorobiphenyls
C12H10
C12H9C1
,*
C12H8C12
C12H9C13
C12H6C14
C12H5C15
C12H4C16.
C12H3C17
C12H2C18
C12H1C19
C12C110
Molecular
weight*
154
188
222
256
290
324
358
392
426
460
490
Percent
chlorine*
0
18.6
31.5
41.0
48.3
54.0
58.7
62.5
65.7
68.5
79.9
No. of
isomers
1
3
12
24
42
46
42
24
12
3
1
*Based on Cl
A-4
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TABLE 2
Physical Properties of Commercial PCBs (Aroclors)
i
Ul
Property
Chlorine, percent
Specific Gravity
Distillation Range
C Corrected
Vapor Pressure
(mm/HS)
Evaporation loss (%)
100 C 6 hr.
USTA D-6 Mod.
160 C, 5 hr.
Pour Point C
(WTM E97) F
1221
20.5-21.5
1.182-1.192
(25°/15.5°C)
275-320
1.0-1.5
1 (Crystal)
34 (Crystal)
1232
31.4-32.5
1.270-1.280
(25°/15.5°C)
290-325
1.0-1.5
-35.5
-32
1016 1242
41 42
1.362-1.372 1.391-1.392
(25°/15.5°C) (25°/15.5°C)
323-356 325-366
4.06xlO~4
0-0.4
3.0-3.6
-19
2
1248
48
1.405-1.415
(65°/15.5°C)
340-375
4.94xlO~4
0-0.3
3.0-4.0
-7
19.4
Water Solubility
at 25 C(ug/l)
>200
225-250
Reference:
Versar, Inc. (1976)
Hammond, et al. (1972)
Hutzinger, et al. (1974)
Mieuer, et al. (1976)
Tucker, et al. (1975)
Mackay and Wolkoff (1973)
240
54
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TABLE 2 (cont)
Physical Properties of Commercial PCBs CAroclors)
Property
Chlorine/ percent
Specific Gravity
Distillation Range
C Corrected
Vapor Pressure
(nun/HS)
Evaporation loss (%)
100 C 6 hr.
USTA D-6 Mod.
160 C/ 5 hr.
Pour Point C
(WTM E97) P
1254
54
1.495-1.555
(65°/15.5°C)
365-390
7.71xlO~5
0-0.2
1.1-1.3
10
50
1260
60
1.555-1.566
(90°/15.5°C)
385-420
4.05xlO~5
0-0.1
0.5-0.8
31
88
1262
61.5-62.5
1.572-1.583
(90°/15.5°C)
390-425
0-0.1
0.5-0.2
35-38
99
1268
68
1.604-1.611
(25°/25°C)
435-450
0-0.6
0.1-0.2
1270
71
1.944-1.960
(25°/25°C)
450-460
Water Solubility
at 25 C(ug/l)
12
2.7
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heating or cooling but at a specific temperature, definesd
as a "pour point," change into a resinous state.
It is known from the studies of pesticides that soil
moisture and evaporation of water have a strong influence
on the rate of chlorinated hydrocarbon volatilization from
soils and sand. Haquef et al (1974) demonstrated that the
periodic evaporization of water from Ottawa sand enhanced
t ' . •
the total volatilization of Aroclor 1254 but reduced the
degree of differentiation in the volatility of the higher
chlorinated biphenyls (7,6, and 5 chlorine aroms) from the
tetrachlorobiphenyls. However, when Aroclor 1254 was heated
in water at 100°C the total volatilization of this Aroclor
was reduced compared to equivalent dry isothermal conditions
but the differentiation in volatility between the higher
and lower chlorinated biphenyls was increased (Bowes, et
al. 1975a) .
Mackay and Wolkoff (1973) calculated theoretical evapora-
tion rates for various Aroclors from water and predicted
very rapid volatilization rates. Under laboratory conditions,
PCBs appear to volatilize fairly rapidly from water in aquaria
(Uhlken, et al. 1973) and even from flasks plugged with
glass wool (Oloffs, et al. 1972). Under the same conditions,
volatilization was markedly reduced in the presence of sedi-
ments (Oloffs, et al. 1973). Hence in natural waters, it
would seem likely that absorption to sediments would limit
the rate of volatilization.
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Solubilities of the individual chlorinated biphenyls
in water have been studied by several workers and an inverse
correlation between solubility and degree of chlorination
has been reported (Wollnofer, et al. 1973; Hague and Schmedding,
1975; Metcalf, et al. 1975). The problem in obtaining true
solution equilibria data for PCBs in water has been explained
by Schoor (1975) who has given evidence that solutions of
PCBs in water are in fact stable emulsions of PCB aggregates
and that the true solubility of Aroclor 1254 is less than
0.1" >ig/l in fresh water and 0.04 jig/1 in marine water.
Chlorobiphenyls are freely soluble in relatively nonpolar
organic solvents (Hutzinger, et al. 1974) and lipids in
biological systems (Hammond, et al. 1972; Metcalf, et al.
1975) . Metcalf, et al. (1975) have reported parition coeffi-
cients between octanol and water in the range of 10,000
to 20,000 for representative tri-, tetra-, and pentachloro-
biphenyls. Partition coefficients with this biphasic solvent
system have been found to correlate well with ecological
magnification factors in aquatic organisms (Metcalf, et
al. 1975).
PCBs are strongly adsorbed on solid surfaces, including
glass and metal surfaces in laboratory apparatus (Schoor,
1975) and soils, sediments, and particulates in the environment
(Haque, et al. 1974; Oloffs, et al. 1973; Crump-Wiesner,
et al. 1974; Dennis, 1976; Munson, et al. 1976; Pfister,
et al. 1969).
In aquatic environments, PCBs are associated with sediments
and are usually found at much higher concentrations in sediments
than in water in contact with them (Young, et al. 1976;
A-8
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Crump-Weisner, et al. 1974; Dennis, 1976). As with other
chlorinated hydrocarbons, PCBs are probably associated partic-
ularly strongly with micro-particulates of 0.15 pm diameter
or less (Pfister, et al. 1969).
PCBs are commercially produced by the chlorination
of the biphenyl ring with anhydrous chlorine in the presence
of iron filings pr ferric chloride as the catalyst. The
crude product is purified to remove color, traces of the
byproduct hydrogen chloride, and the catalyst by treatment
with alkali and subsequent distillation. The purified product
is a complex mixture of the chlorobiphenyls, the precise
composition depending on the conditions under which the
chlorination occurred.
It has been reported that foreign PCB mixtures are
similar in composition to one of the 10 Aroclor products
previously manufactured in the U.S. Gas liquid chromatograms
of Phenoclor DP6 (France), Clopen A60 (Germany), and Aroclor
1260 (U.S.), all mixtures containing 60 percent chlorine,
Show that these mixtures are virtually identical (Tas. and
de Vos, 1971). Jensen and Sundstron (1974) have shown that
Clophen A60 and A50 (Germany) are very similar in isomer
composition to Aroclors 1260 and 1254 (U.S.) respectively.
„Table 3 lists the distribution of the various classes of chloro
biphenyls in seven major Aroclor mixtures as reported by
Mieure, et al. (1976) Webb and McCall (1973), and Hirwe,
et al. (1974).. The,small differences in analytical results
reported for Aroclors 1242 and 1254 may reflect either dif-
ferences in analytical methods or variations in sample constitu
tion.
A-9
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TABLE 3
Approximate Molecular Composition of A'roclors
Chlorobiphenyl
1221 1232 1242 1248 1254 1261
M W W M M W H W M W H W
C12H
C12H10C1
C12H9G12
C12H8C13
G12H7G14
C12H6C15
C12H5G16
GI2H4C17
11 7 6 Tr Tr Tr
51 51 26 1 1 • 1 Tr . - Tr
32 38 29 20 16 17 4 1 0.5 -
4 3 24 57 49 40 39 23 1 - 0.5
2 - 15 21 25 32 42 50 21 16 36
0.5 0.5 1 8 10 14 20* 48 60 45 12
- Tr 1 0.5 1 23 23 18' 46
- - Tr - - 61 1 36
Tr - Trace (less than Ovl percent) Letters refer to references
Reference: Micure, et al. (1976)
Webb and McCall (1973)
Hirwe, et al. (1974)
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Certain substitution patterns are believed to influence
the biological activities of chlorobiphenyls. The presence
of two adjacent carbon atoms without chlorine substitution
in one or both rings is believed to facilitate metabolism
because it permits the formation of arene oxide intermediates
(Safe, et al. 1975). Essentially all chlorobiphenyls with
five or fewer chlorine atoms have at least one pair of adjacent
unsubstituted carbon atoms because of the rarity of 3,5-
substitution in the natural mixtures.
Jensen and Sundstrom (1974b) presented evidence that
chlorobiphenyls with three or four chlorine atoms in the
ortho-positions (2- and 6- positions) are more easily meta-
bolized by humans than those with only one or two ortho-
chlorines. Compounds with three or four ortho-substituted
chlorines are virtually absent from Aroclors 1016 and 1242
but are fairly well represented in Aroclors 1254 and 1260
(Clopens A50 and A60 respectively).
McKinney (1976) has suggested that chlorobiphenyl isomers
with chlorine substitution in both the 4- and 41 positions
tend to be biologically active and well retained in tissues.
The number and proportion of these isomers increase with
increasing mixture chlorination.
McKinney, et al. (1976a) have shown an association
.between biological activity and substitutions in the 3,4-,
or 3,4,5- positions on one or both rings. The first pattern
is frequently found in PCB mixtures but the second is found
only as part of the 2,3,4,5-pattern which is found in only
trace amounts in PCBs.
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Toxic materials other than chlorinated biphenyls have
been found in commercial PCS mixtures. Vos, et al. (1970);
Bowes, et al. (1975a), Roach and Pomerantz (1974), Nagayama,
et al. (1976), and Kuratsume, et al. (1976) have detected
polychlorinated dibenzofurans (PCDFs) in a number of domestic
and foreign PCB mixtures at levels of 0.8 to 33 mg/kg.
While 119 structurally different PCDF isomers are possible,
only two have been precisely identified to date, the 2,3;7;8-
tetrachloro- and the 2,3,4,7,8-pentachlorodibenzofurans
(Bowes, et al. 1975).
Polychlorinated naphthalenes (PCNs) have also been
identified in small quantities in Clopen A60 and Phenochlor
DP 6 (both corresponding to Aroclor 1260), Aroclor 1254>
and KC-400 (corresponding to Aroclor 1248) (Vos, et al.
1970; Roach and Pomerantz, 1974; Bowes, et al. 1975).
There appear to be no authenticated reports of poly-
chlorinated dibenzo-p-dioxins (PCDDs) in commercial PCBs
(Bowes, et al. 1975a). The presence of potentially toxic
compounds other than polychlorinated biphenyls in commercial
PCB mixtures complicates both analytical and toxicological
evaluation of such mixtures.
PCBs are considered to be inert to almost all of the
typical chemical reactions. PCBs do not undergo oxidation,
reduction, addition, elimination, or electrophilic substitu-
tion reactions except under extreme conditions. Chlorines
can be replaced by reductive dechlorination with any metal
hydride such as lithium aluminum hydride but temperatures
of 245°C or greater are required to effect chlorine displacement;
A-12
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The reactions of environmental importance that PCBs
appear to undergo include alkali- and photochemically-catalized
nuceleophilic substitutions and photochemical free radical
substitutions, all of which occur with alkali and water.
Photolysis generally has been found to give one type
of product under environmental conditions (Hutzinger, et
al. 1974; Ruzo, et al. 1972; Ruzo, et al. 1974a; Ruzo and
Zabik, 1975; Hutzinger, et al. 1972c; Herring, et al. 1972).
Chlorine is replaced by hydroxy groups in aqueous systems.
A marked increase in rate of PCB photolysis was observed
when solvents were degassed prior to irradiation (Ruzo,
et al. 1974a). Oxygen is known to act as a free radical
quencher by accepting energy from free radicals before any
chemical change can occur. This increase in rate therefore
implies that a free radical process is occurring and in
the environment these photochemical transformations will
be enhanced under anaerobic conditions.
The photochemical behavior of higher chlorobiphenyls
appears similar to that of the tetrachlorobiphenyls (Hutzinge'r,
et al. 1972c; Herring, et al. 1972). Irradiation of Aroclor
1254 in aqueous solution gave rise to dechlorinated and
hydroxylated products (Hutzinger, et al. 1972c). Hexa- and
octachlorobiphenyls are more photochemically reactive than
tetrachlorobiphenyls (Hutzinger, et al. 1972c), so that
under irradiation the higher components of Aroclor 1254
are selectively degraded (Hutzinger, et al. 1972c; Herring,
et al. 1972).
A-13
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The creation of free radicals by sunlight allows the
environmental replacement of chlorines by hydroxy groups
from water without the intervention of alkali. When this
occurs at the ortho position (found to the the most preferred
for chlorine loss) the resulting 2-hydroxychlorobiphenyl
is perfectly positioned to allow oxygen to bond to an ortho
position of the other ring. This results in the creation
of potentially the most important class of contaminant in
commercial mixtures of PCBs, the chlorodibenzofurans (CDFs).
Irradiation studies on either Aroclor 1254 or 2,5,2', 51-
tetrachlorpbiphenyl (Hutzinger, et al. 1972c) in hydroxylic
solvents have shown the formation of phenolic compounds,
carboxylic compounds, and polymers along with dechlorination.
Activation of the phenyl rings by metals or metallic salts
make them more susceptible to hydroxylation. Thus in the
environment, either heat, light, or metals and metal salts
in water could theoretically accelerate the transformation
of PCBs to PCDFs. The ultraviolet component of sunlight
is sufficiently energetic to generate free radicals from
both phenols and PCBs. The energies required to break the
Ar-Cl bond to form hydroxy-PCBs in a hydroxylic solvent
and ArO-H bond to form CDFs correspond to wavelengths near
360 to 320 nm, respectively. These wavelengths are clearly
within the sunlight region.
Irradiation experiments with five pure 2-chlorinated
biphenyls as 5 mg/1 aqueous suspensions, showed that traces
of 2-chlorodibenzofuran were detectable although only the
•2,5-dichloro- and the 2,5,2',5'-tetrachlorobiphenyls provided
identifiable amounts or approximately a 0.2 percent yield
A-14
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during a seven-day irradiation (Crosby, et al. 1973; Crosby
and Moilanen, 1973). The environmental significance of
this is four fold: (1) ortho-chlorobiphenyls can be hydroxy-
lated by radiation similar to sunlight when they are suspended
in aqueous media; (2) the product(s) are converted to CDFs;
(3) rates of CDF formation by this process are approximately
the same as their rates of degradation, leading to an approxi-
mately steady concentration. The fourth point of significance
is illustrated by irradiation studies on 2,8-dichlorobenzo-
furan (Crosby and Moilanen, 1973). Decomposition of this
material was found to be very slow in aqueous suspension
but dehalogenation did not take place to form the relatively
photdlytically stable 2-chlorodibenzofuran.
In addition to photochemical and metallic/metallic
salt formations of PCDFs from PCBs, a third route of formation
•*
has been suggested. Kanechlor KC-400 (analogous to Aroclor
•» i
1248) having an intitial PCDF content of 20 mg/kg, was shown
to undergo conversion as the heat transfer fluid in a heat
exchanger to give PCBs with a PCDF content of 4975-11765
mg/kg '(Nagayma, et al. 1976; Kuratsune, et al. 1976). This
V
material was identified as the agent which poisoned a large
number of Japanese in 1968. A general disadvantage of PCBs
in many of their applications including electrical capacitor
'.*' *•
and transformer uses as well as heat transfer uses is their
tendency to decompose under the action of heat or electrical
arcing to form potentially more toxic products (Broadhurst,
A-15
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'
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A-16
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A-17
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.. i
•' •«
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A-18
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.'
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s •
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A-19
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major components in a technical polychlorinated biphenyl
mixture. Environ. Sci. Tech. 5: 1216.
Tucker, E.S., et al. 1975. Migration of polychlorinated
biphenyls in soil induced by percolating water. Bull. Environ,
Contain. Toxicol. 13: 86.
Uhlken, L.D., et al. 1973. Apparent volatility of PCB's
as used in continuous flow bioassays. PCB Newsletter 5: 4.
Versar, Inc. 1976. Final Report. PCBs in the United States:
Industrial use ^id environmental distribution. Report to
U.S. Environmental Protection Agency. Task I: Contract
No. 68-01-3259.
Vos, J.G., and J.H. Koeman. 1970. Comparative toxicologic
study with polychlorinated biphenyls in chickens with special
reference to porphyria, edema formation, liver necrosis,
and tissue residues. Toxicol. Appl. Pharm. 17: 656.
Webb, R.G., and A.C. McCall. 1973. Quantitative PCB standards
for electron capture gas chromatography. Jour. Chromatog.
Sci. 11: 366.
A-20
-------
Wollnofer, P.R., et al. 1973. The solubilities of twenty-
one chlorobiphenyls in water. Analabs Research Notes 13: 14
Young, D.R., et al. 1976. Marine inputs of polychlorinated
biphenyls off southern California. Proc. Nat. Conf. on
Polychlorinated Biphenyls: 197-209.
A-21
-------
AQUATIC LIFE TOXICOLOGY*
FRESHWATER ORGANISMS
Introduction
Most data for polychlorinated biphenyls (PCB's) are for
studies concerned with tissue levels in fish, mammals, and birds,
without correlation with source or exposure concentrations. Many
studies dealing with various physiological parameters are also
available, but again, are such that they are of little use here..
Also, PCB's often do not appear to be very acutely toxic to
juvenile and adult freshwater fish and invertebrate species due to
solubility problems in static tests, and this can lead to
erroneous judgments as to the actual toxicity of the compounds.
PGB.'s occur as mixtures of chemical isomers that differ in
the amount of chlorination of the biphenyl structure, they have
been treated herein as a single entity. They are highly
lipophilic and bioconcentrate to high tissue concentrations from
*The reader is referred to the Guidelines for Deriving Water
Quality Criteria for the Protection of Aquatic Life [43 FR 21506
(May 18, 1978) and 43 FR 29028 (July 5, 1978)] in order to better
understand the following discussion and recommendation. The
following tables contain the appropriate data that were found in
the literature, and at the bottom of each table are the calcula-
tions: for deriving various measures of toxicity as described iVi
.the Guidelines.
B-l
-------
water concentrations that are below the usual detection limits*
Acute Toxicity
Only three 96-hour LC50 values are available and these are
for the fathead minnow (Table 1). Newly-hatched fish were more.
sensitive than juveniles with LC50 values of 15 ug/1 and 300 ug/1,
respectively, for Aroclor 1242 (A-1242). A-1254 was more toxic
with an LC50 of 7.7 ug/1 for newly-hatched fish. No adjustments
to the data were necessary as all values were from flow-through
tests with measured concentrations. Other 96-hour LC50 values in
the literature exceeded the solubility of PCB's and were not used.
The solubility of PCB's is, at most, 250 ug/1 and 96-hour LC50
values such as 50,000 ug/1 are meaningless.
The acute toxicity data base for invertebrate species (Table
2) contains 12 values for three species. These values were from
static and flow-through tests and showed an LC50 range from 10
ug/1 for the scud, Garamarus fasciatus, to 200 to 400 ug/1 for the
damselfly, Ischnura verticalis. The higher chlorinated isomers,
such as Aroclor 1254 which contains 54 percent chlorine, were more
toxic to Gammarus pseudolimnaeus, than they were to fish, with
LC50 values of 73 ug/1;for A-1242 and 29 ug/1 for A-1248.
Acute toxicity tests with polychlorinated biphenyls have
established that these compounds can be toxic to aquatic life at
low concentrations once they are in solution. The data indicate
that the more highly chlorinated compounds are more toxic to fish
and invertebrate species.: The Final Fish Acute Value is 7.7 ug/1
and the Final Invertebrate Acute Value is 6.2 ug/1 which also
B-2
-------
becomes the Final Acute Value.
Chronic Toxicity
Four flow-through chronic tests have been conducted with
polychlorinated biphenyls and fathead minnows. Test
concentrations were measured. The fish were most sensitive to
A-1248 with a chronic value of-0.2 ug/1 (Table 3). Chronic values
for A-1242, A-1254, and A-1260, were 9.0, 2.9, and 2.3 ug/1,
respectively. The geometric mean (1.86 ug/1) of the four chro(nic
concentrations divided by the species sensitivity factor (6.7)
results in a value of 0.28 ug/1. Since the lowest chronic value
is 0.20 ug/1, it becomes the Final Fish Chronic Value.
Results from 14 chronic tests with 2 invertebrate species,
Daphnia magna and Gammarus pseudolimnaeus, are shown in Table 4.
The low chronic values for Daphnia magna, 1.14 ug/1 for A-1242,
1.4 ug/1 for A-1248, and 0.73 ug/1 for A-1254 were from flow-
through tests with measured concentrations; the other values were
from static tests, and were much higher due to loss of PCB's from
the test containers. The two chronic values for Gammarus
pseudolimnaeus, 4.9 ug/1 for A-1242 and 3.3 ug/1 for A-1248, were
from flow-through tests with measured concentrations. The
geometric mean of the 14 tests was 8.1 ug/1 which, after division
by the species sensitivity factor (5.1) results in a concentration
of 1.6 ug/1. Since there is a lower chronic value (0.73 ug/1 for
Daphnia magna), this latter value becomes the Final Invertebrate
Chronic Value.
B-3
-------
Plant Effects
Results from six tests with four different algal species
are shown in Table 5. In general the data show that plants were
less sensitive than the fish and invertebrate species, but
reduction in the rate of carbon fixation in Scenedesmus
quadricauda occurred at 0.1 ug/1 A-1254 which is lower than the
Final Fish and Invertebrate Chronic Value. Therefore, the Final
Plant Value is 0.1 ug/1.
i
Residues
Table 6 contains results of 25 laboratory residue studies and
47 field studies of fish residues where information on water
concentrations was also available. The studies include laboratory
and field data for invertebrate and fish species and show a wide
range of bioconcentration factors (BCF's). For invertebrate
species the BCF ranged from 740 for stoneflies exposed for 21 days
to 125,000 for mysids collected from Lake Superior. The BCF for
fish ranged from 3,500 for field collected bass to 4,125,000 for
field-collected siscdwet, a race of lake trout. For laboratory
exposures, the BCF ranged from 5,500 for white sucker exposed for
30 days to 540,000 for minnows exposed for 240 days.
The residue limit established by the Food and Drug
Administration (FDA) for polychlorinated biphenyls in edible fish
and shellfish is 5.0 mg/kg. Significant effects on reproduction
oE mink were observed when fed food containing 0.64 mg/kg; this
figure was used to calculate the Residue Limited Toxicant
B-4
-------
Concentration (RLTC). Since fish is one of the principal foods of
mink, the mink-effect concentration of 0.64 mg/kg was divided by
the geometric mean fish bioooncentration factor of 427,000 to give
an RLTC of 0.0000015 mg/kg or 0.0015 ug/1.
The lowest of the .Final Fish Chronic Value (0.2 ug/1), Final
Invertebrate Chronic Value (0.73 p.g/1) , Final Plant Value (0.1
ug/1) and the RLTC (0.0015 ug/1) is used to determine the Final
Chronic Value. For polychlorinated biphenyls the Final Chronic
Value is 0.0015 ug/1.
Miscellaneous
Data presented in Table 7 do not conflict with the selection
of 0.0015 ug/1 as the Final Chronic Value.
B-5
-------
Criterion Formulation
Freshwater Aquatic Life
Summary of Available Data
Final Fish Acute Value + 7.7 pg/1
Final Invertebrate Acute Value = 6.2 pg/1
Final Acute Value = 6.2 pg/1
Final Fish Chronic Value =0.20 pg/1
Final Invertebrate Chronic Value = 0.73 pg/1
Final Plant Value = 0.10 pg/1
Residue Limited Toxicant Concentration = 0.0015 pg/1
. Final Chronic Value = 0.0015 pg/1
0.44 x Final Acute Value = 2.7 pg/1
The maximum concentration of polychlorinated biphenyls
is the Final Acute Value of 6.2 pg/1 and the 24-hour average
concentration is the Final Chronic Value of 0.0015 pg/1.
'' \
No important adverse effects on freshwater aquatic organisms
have been reported to; be caused by concentrations lower
than the 24-hour average concentration.
CRITERION: For polychlorinated biphenyls the criterion
to protect freshwater aquatic life as derived using the
Guidelines is 0.0015 pg/1 as a 24-hour average and the concen-
trations should not exceed 6.2 pg/1 at any time.
B-6
-------
Table 1. Freshwater fish acute values for polychlorinated biphenyls (Nebeker. et al. 1974)
03
I
Organism
Fathead minnow
(juvenile),
Pimephales promelas
Fathead minnow
(newly hatched) ,
Pimephales promelas
Fathead minnow
(newly hatched) ,
Pimephales promelas
Bioassay Test • .
Mfctnod* Cone .**
FT M
FT M
FT M
Chemical
Description
A-1242
A-1242
A-1254
Adjusted
Time LC&u LCio
(hrs) (uq/JL) (uq/i)
96 300 300
t
96 15 15
96 7.7 7.7
* FT = flow-through
*.* M = measured
Geometric mean of adjusted values - 33 vg/1 39 =8.4 pg/1
Lowest value from a flow-through test with measured concentrations =7.7 yg/1
33
-------
Table 2. Freshwater invertebrate acute values for polychlorinated biphenyls
09
I
oo
Or nanism
Scud.
Gammarus
Scud,
Gammarus
Scud,
Gammarus
Scud,
Gammarus
Scud,
Gammarus
Scud,
Gammarus
Scud,
Gammarus
Scud,,
Gammarus
Scud,
Gammarus
Scud,
Gammarus
Bioassay
Method*
..-..-. .FT
fasciatus '-•••.-;
fasciatus
fasciatus
pseudolinmaeus
p s eudo 1 imnaeus
pseudolinmaeus
pseudolinmaeus
pseudolinmaeus
pseudolinmaeus
pseudolimnaeus
Damselfly,
Ischnura verticalis
Damselfly,
Ischnura verticalis
S
S
FT
FT
S
S
S
S
S
FT
FT
Test
Cone. *
M
U
U
M
M
U
U
U
U
U
M
M'
Chemical Time
* . Description (hrs)
A-1242 96
A-1248 •
A-1254
A-1242
A-1248
2.3,4- tri-
chlorobiphenyl
4,4 - dichloro-
biphenyl
2,4- dichloro-
biphenyl
2. 4, 6,2', 4', 6-
hexachloro-
biphenyl
2452' 5 -
£, H, j , £ , j -
pentachloro-
biphenyl
A-1242
A-1254
96
96
96
96
96
96
96
96
96
96
96
Adjusted
LCbo LCiu
(gq/1^ (ua/1)
10
52 •',:.;.'•
2,400
73
29
70
100
120
150
210'
400
200
10
44
2,032
73
29
59
85
101
127
178
400
200
Reference
Mayer ,
Mayer,
Mayer,
Nebeker
1974
Nebeker
1974
Mayer,
Mayer.
Mayer,
Mayer,
Mayer,
Mayer ,
Mayer,
et
et
et
&
&
et
et
et
et
et
et
et
al.
al.
al.
1977
1977
1977
Puglisi.
Puglisi,
al. 1977
al.
al.
al.
al.
al.
al.
1977
1977
1977
1977
1977
1977
* S =• static, FT » flow-through
** U = unmeasured, M = measured
Geometric mean of adjusted values = 131- ng/1 21 - 6.2 jig/1
Lowest value from a flow-through test with measured concentrations
131
10 pg/1
-------
Tacle 3. Freshwater fish chronic values for poiychlorinated biphenyls
Chronic
Limits Value
Organism
Fathead minnow,
Pimephales promelas
Fathead minnow,
Pimephales promelas
Fathead minnow,
Pimephales promelas
Fathead minnow,.
Pimephales promelas
Test*
LC
LC
LC
LC
taq/H
0.1 - 0.4
A- 1248
1.3 - 4.0
A- 1260
5.4 - 15
A-1242
1.8 - 4.6
A-1254
(uq/i)
0.2
2.3
9.0
2.9
Reference
DeFoe, et al, .In
DeFoe, et al. in
Nebeker, et al.
Nebeker, et al.
press
press
1974
1974
* LC » Life cycle or partial life cycle
1.86
I Geometric mean of chronic values = 1.86 ug/1 6.7 = 0.28 wg/1
vo
Lowest chronic value =0.2 pg/1
Application Factor Values (Nebeker, et al. 1974)
96-hr LC50 MATC
Species (pg/1) (ug/1) AF
Fathead minnow, 15.0 '9.0 0.6
Pimephales promelas (A-1242) (A-1242)
Fathead minnow, 7.7 2.9 0.38
Pimephales promelas (A-1254) (A-1254)
Geometric mean AF - 0.48 Geometric mean LC50 = 10.75 pg/1
0.48 S7.7 pg/1 x 10.75 pg/1 = 4.4 pg/1
-------
Table 4. Freshwater invertebrate chronic values for polychlorinated biphenyls
CO
M
O
Organism
Cladoceran,
Daphnia m'agna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran.
Daphnia magna
Cladoceran.
Daphnia magna
Cladoceran,
Daphnia magna
Scud,
Gamma r us pseudolimnaeus
'Scud,
Cammarus .pseudolimnaeus
Test*
LC
LC
LC
LC
.LC
LC
LC
LC
LC
LC
LC
LC
LC
LC
Limits
10-24
89-125
53-66
48-63
16-24
18-28
22-33
24-41
162-206
1.0-2.1
0.48-1.1
1.0-1.3
2.8-8.7
2.2-5.1
Chronic
Value
(uq/1)
15 '" "''•••
A-1254
105 '
A-1221
59
A-1232
55
A-1242
19
A-1248
22
A-1254
27
A-1260
31
A-1262
182
A-1268
1.4
A-1248
0.73
A-1254
1.14
A-1242
.4.9
A-1242
3.3
A-1248
Reference
Maki & Johnson, 19 75-
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
Nebeker
&
&
&
&
&
&
&
&
&
&
&
&
&
Puglisi ,
Puglisi,
Puglisi.
Puglisi.
Puglisi.
Puglisi.
Puglisi.
Puglisi.
Puglisi.
Puglisi,
Puglisi,
Puglisi,
Puglisi,
1974
1974
1974
1974
1974
,1974
1974
1974
•1974
.1974
1974
.1974
1974
8.1
* LC = Life cycle or "partial life cycle
Geometric mean of chronic values -8.1. pg/1 « 1.6 ug/1
Lowest chronic value - 0.73
-------
Table 5. Freshwater plant effects for polychlorinaced biphenyls
Organism
Effect
Cone e n tr at i on
(ug/j.)
Reference
Alga.''
Chlamydompnas
reinhardt-ii
Reduced
growth
2.000
A-1242
Morgan, 1972
BJ
I
Alga.
Chlorella
pyrenoidosa
Alga,
Chlorella
pyrenotdosa
Alga,
Euglena gracilis
Alga, .
Scenedesmus
obtusiusculus
Alga,
Scenedesmus
quadricauda
Depressed 1,000
cell produc- A-1268
tivlty
Reduced 1,000
population A-1254
growth
48 hr 4,400
ID50 A-1221
Growth 300
inhibition A-1242
Reduction in 0.1
rate of carbon A-1254
fixation
Hawes, et al. 1976b
Hawes, et al. 1976a
Ewald, et al. 1976
Larsson & Tillberg, 1975
Luard, 1973
Lowest plant value = 0.1 vig/1
-------
Table 6. Freshwater residues for polychlorinated biphenyls
Organism
Snail.
Physa sp.
Snails,
Cladoceran,
Daphnia magna
Mysid,
Mysis relicta
Scud,
Garamarus pseudolimnaeus
Scud,
Gamma rus pseudolinmaeus
Amphipod ,
W Pohtiporeia affinis
1
lyj Glass shrimp,
Palaemonetes kadiakensis
Crayfish,
Orconectes nais
Stonefly,
Pteronarcys dorsata
Mosquito,
Culex tarsalis
Phantom midge,
Chaoborus punctipennta
Dobsonf ly ,
Corydalus cornutus
Gizzard shad.
Dorosoma cepedianum
Alewife,
Bioconcentration factor
59,600
45,000
3,800
125,000
6.200
108,000
1,709
2,600
750
740
3.500
2,700
1.500
150,300
270,000
Time
(days)
33
Field data
4
Field data
21
60
Field data
21
21
21
7
14
7
Field data
Field data
Reference
Sanborn, 1974
Nadeau &.\f)avis
Mayer, et al.
. 1976
1977
Veith, et al. 1977
Mayer, et al. 1977
Nebeker & Buglisi, 1974
Halle, et al.
Mayer, et al.
Mayer , et al .
Mayer, et al.
Mayer, et al.
Mayer, et al.
Mayer, et al.
Hesse, 1973
Hesse. 1973
1975
1977
1977
1977
1977
1977
1977
Alosa pseudoharengus
-------
Table 6. (Continued)
Organism
Bioconcentratidn Factor
Time
(days) Reference
Alewife,
Alosa pseudoharengus
Alewife,
Alosa pseudoharengus
Chub,
Coregonus j oh anna e
Me nominee,
Pros opium cylindraceum
Lake whitefish,
Coregonus clupeaformis
Lake whitefish,
Coregonus clupeaformis
CD
jL Bloater,
oj Coregonus hoyi
Bloater,
Coregonus hpyi
Lake herring,
Coregonus artedii
Rainbow trout,
Salmo gairdneri
Rainbow trout ,
Salmo gairdneri
Rainbow trout,
Salmo gairdneri
Steelhead trout,
Salmo gairdneri
Brook trout,
Salvelinus fontinalis
Brook trout,
89,000
42,700
850,000
120,000
110., 000
875,000
1,162,500
81,000
250,000
120,000
46,000
5,850
600,000
47 , 000
60,000
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
30
42
Field data
118
500
Veith, 1975'
Haile, et al. 1975
Hesse, 1973
Hesse, 1973
Hesse, 1973
Veith, et al. 1977
Veith, et al. 1977
Veith, et al. 1977
Veith. et al. 1977
Veith, 1975
Bills & Marking, 1977
Branson, et al. 1975
Hesse, 1973
Mauck, et al. In presi
Snarski & Puglisi, 19
Salvelinus fontinalis
-------
Table 6. (Continued)
a
i
M
*>.
Organism
Brown trout,
Salmo trutta
Lake trout,
Salvelinus namaycush
Lake trout,
Salvelinus namaycush
Lake trout,
Salvelinus namaycush
Lake trout.
Salvelinus namaycush
Siscowet,
S. namaycush siscowet
Chinook salmon,
Oncorhynchus tschawytscha
Chinook salmon.
Oncorhynchus tschawytscha
Coho salmon,
Oncorhynchus kisutch
Rainbow smelt,
Osmerus mordax
Rainbow smelt,
Osmerus mordax
Rainbow smelt,
Osmerus mordax
Pike,
Esox lucius
•— — .^— *^» •^••H^W^^^^B^to
Carp,
Cyprinus carp to
Carp,
Bioconcentration Factor
119,000
1,110,000
212,000
2,333,000
1,625,000
4,125,000
1,240,000
240,000
173,000
462,500
32,000
48,000
15,000
43,600
390,000
Time
(days)
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Reference
Veith, 1975
Hesse, 1973
Veith, 1975
Parejko, et al.
Veith, et al.
Veith, et al.
Veith, et al.
Hesse, 1973
Veith, 1975
Veith, 1975
Veith, et al.
Veith. 1975
Haile, et al.
Hesse. 1973
Hesse, 1973
Hesse, 1973
. 1975
1977
1977
1977
1977
1975
Cyprinus'carpio
-------
Table 6. (Continued)
Organism
'
Carp,
Cyprinus carpio
Fathead minnow.
Pimephales promelas
Fathead minnow,
Pimephales promelas
Fathead minnow,
Pimephales promelas
Fathead mtnow,
Pimephales promelas
Fathead minnow,
Pimephales promelas
03 *
t_i Fathead minnow.
tn Pimephales promelas
Fathead minnow,
Pimephales promelas
Fathead minnow.
Pimephales promelas
Common shiner,
Notropis cornutus
Longnose sucker,
Catostomus catostomus
Longnose sucker,
Catostomus catostomus
Redhorse sucker,
Moxostoma sp.
White sucker,
Catostomus cotnmersoni
Bioconcentration Factor
110,000
240,000
120,000
270,000
540,000
274,000
107,000
235,000 .
238,000
> 78, 000
150,000
1,125,000
32,000
106,000
Time
(days)
• • v
Field data
240
240
240
240
255
255
240
240
Field data
Field data
Field data
Field data
Field data
Reference
";; '
Veith, 1975
DeFoe, et al.
Defoe, et al.
DeFoe, et al.
DeFoe, et al.
Nebeker, et al
Nebeker, et al
Nebeker, et al
Nebeker, et al
Nadeau & Davis
Hesse, 1973
Veith, et al.
Veith, 1975
Veith, 1975
In 'press
la press
In 'press
In press
. 1974
. 1974
. 1974
. 1974
, 1976
1977
White sucker,
Catostomus commersoni
5,500
30
Frederick, 1975
-------
Table 6. (Continued)
09
Organism
Bioconcentration Factor
Time
(days) Reference
Channel catfish,
•Ictalurus punctatus
Burbot,
Lota lota
Rock bass,
Ambloplites rupestris
Bluegill.
Lepomis macrochirus
Largemouth bass ,
Micropterus salmoides
Yellow perch,
Perca flavescens
Yellow perch,
Perca flavescens
Yellow perch,
Perca flavescens
Yellow perch,
Perca flavescens
Slimy sculpin.
Cottus cognatus
Slimy sculpin,
Cottus cognatus
Fourhorn sculpin,
49,000
1,162,500
117,000
52,000
3,500
14,800
50.000
109.000
154,000
300,000
84,000
337.500
, ^hMH^H^^H^-M^B*
77
Field data
«' ^
Field data
77
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Field data
Stalling. 1971
Veith, et al.
Nadeau & Davis ,
Stalling. 1971
Martell, et al.
Hesse. 1973
Hesse, 1973
Veith, 1975
Nor strora, et al
Veith. et al.
Haile, et al.
Veith, et al.
1977
1976
1975
. 1976
1977
1975
1977
Myoxocephalus quadricornis
-------
Table 6. (Continued)
Maximum Permissible Tissue Concentration
Organism
Man
Mink.
Mustela vison
Mink,
Mustela vison
Concentration
Action Level or Effect (mg/kg)
*
Edible fish and shellfish 5
FDA action level
Reduced reproduction 1
No- reproduction, mortality 0.64
Reference
21 CFR Part 122.10
Ringer, et al. 1972
Platonow & Karstad, 1973
CO
I
Geometric mean fish bioconcentration factor = 427,000
Lowest residue concentration = 0.64 mg/kg
42°;000 • 0.0000015 mg/kg or 0.0015 yg/1
-------
Table 7. Other freshwater data for polychlorinated biphenyls
Organism
Cladoceran,
Daphnia pulex
Cladoceran, „"".'
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Ostracod,
Cypridopsis vidua
03
!_, Glass shrimp,
oo Palaemonetes kadiakensis
Crayfish,
Orconectes nais
Dragonfly ,
Macromia sp .
Midge,
Tanytarsus dissimilis
Mosquito.,
Culex tarsalis
Rainbow trout,
Salmo gairdneri
Rainbow trout.
Salmo gairdneri
'Rainbow trout.
Salmo gairdneri
Rainbow trout.
Salmo gairdneri
Test
Duration
4
days
2 •-,
wks
2
wks
3
wks
2
wks
3
days
7
days
7
days
7
days
3
wks
7
days
—
25
days
25
days
25
days
Ett'ect
Significant
mortality
LC50
LC50
LC50
LC50
Significant
mortality
LC50
LC50
LC50
50% death
of pupae
No adult
emergence
Inhibit ATPase
activity
LC50
LC50
LC50
Result
tuq/1)
2,000
A-1242
2.6
A-1248
1.8
A-U54
1.3
A-1254
24
A-1254
2,000
A-1242
3
A-1254
30
A-1242
800
A-1242
0.45
A-1254
1.5
A-1254
4 pg/g
A-1242
12
A-1242
3-4
A-1248
27
A-1254
Reference
Morgan, 1972
Nebeker & Puglisi.. 1974
Nebeker & Puglisi. 1974
Nebeker & Puglisi, 1974
Maki & Johnson, 1975
Morgan, 1972
Mayer, et al. 1977
Mayer, et al. 1977
Mayer, etal.. 1977
Nebeker£ Puglisi, 1974
Sanders & Chandler. 1972
Davis, et al. 1972
Mayer, et ,al. 1977
Mayer, et al. 1977
Mayer, et al. 197.7
-------
Table 7. (Continued)
Organism
Test
Duration Etfect
flesult
Rainbow trout,
Salmo gairdneri
Rainbow trout,
Salmo gairdneri
Rainbow trout,
Salmo gairdneri
Rainbow trout ,
Salmo gairdneri
Rainbow trout,
Salmo gairdneri
QJ Rainbow trout,
1 Salmo gairdneri
*° Steelhead trout,
Salmo gairdneri
Coho salmon,
Oncorhynchus kisutch
Coho salmon,
Oncorhynchus kisutch
Coho salmon,
Oncorhynchus kisutch
V
Coho salmon,
Oncorhynchus kisutch
Atlantic salmon,
Salmo salar
Atlantic salmon,
Salmo salar
25
days
21
days
30
days
330
days
5
days
5
days
24
days
embryo-
larval
72
hrs
68
days
72
days
96
hrs
192
hrs
.1 i • • — T ••
LC50
Induce fish.
hepatic micro-
somal enzymes
75% mortality. '70%
deformed fry
Kidney pathology
LC50
LC50
Bioconcentration
factor
MATC
Stimulated
thyroid
activity
Induced fish
hepatic AHH
microsomal
enzymes
Induction of aryl
hydrocarbon
hydroxylase
Bioconcentration
factor
Mortality
mm m -J • mm
49
A-1260
31 yg/g "'
Clophen
A-50
0.39 yg/g
PCB
10 yg/g
A-1254
67
A- 1242
54
A-1254
38.000
times
<4.4
0.48 yg/g
A-1254
1 ng/g
A-1242
12 yg/g
PCBYs
600
times
>2 yg/g
A-1254
Mayer,
Lidman
et al.
, et al.
Hogan & Brauhn.
Nestel
Mayer ,
Mayer,
Halter
Halter
Mayer ,
Gruger
Gruger
Zitko
Zitko,
& Budd,
et al.
et al.
. 1974
1977
1976
1975
1974
1977
1977
& Johnson, 1974
et al. 1977
, et al.
, et al.
& Carlson
1970
1977
1976
, 1977
-------
Table 7. (Continued)
Organism
Brook trout,
Salvel-inus fontinalis
Brook trout,
Salvelinus fontinalis
Brook trout,
Salvelinus fontinalis
Brook trout,
Salvelinus fontinalis
Brook trout,
Salvelinus fontinalis
03
1 Brown trout,
ro Salmo trutta
o
Northern pike (fry).
Esox lucius
Carp.
Cyprinus carpio
Carp,
Cyprinus carpio
Fathead minnow0,
Pimephales promelas
Fathead minnow, .
Pimephales promelas
Fathead minnow.
Pimephales promelas
Test
Duration
embryo-
. larval
71
wks
fert. .-.-
to
hatch
21
days
21
days
43
days
field
data
20
days
21
days
30
days
• 30 ;?••
days
30
days
Ettect
Result
(uq/ll
MATC <0.43
.V" • ' •
No effect on survival 0.94
growth or reproduction A-1254
No egg hatch 200 .
A-1254
Stimulated 200
hydroxylation A-1254
of testosterone
Bioconcentration 164
factor times .
Anaemia 10 pg/g in
hyperglycaemia food Clophen
altered cholestrol A-50
metabolism
Possible
mortality
Altered plasma
0-glucoronidase
activity
Hjtabolic
changes
30- day LC50
';• - •
Reduced growth
1.41 yg/g
tissue
1.8 pg/g
eggs
A-1248
5 yg/g
A-1248
250 pg/g
A-1248
28
A-1242
28
A- 1016
23
A- 1016
RetereiiCfe
Mauck, et al," In
Snarski & Puglisi
Freeman & Idler,
Freeman & Idler,
Freeman & Idler,
Johansson, et al
Waybrant. 1974
ltd. 1973
press
.1976
1975
1975
1975
1972
Ito & Murata, 1974
Veith. 1976
Veith. 1976
Veith. 1976
-------
Table 7. (Continued)
Organism
Test
Duration Effect
Result
(ug/l) ... Reference
03
I
to
Fathead minnow, 30
Pimephales promelas days
Fathead minnow, 30
Pimephales promelas days
Fathead minnow, 30
Pimephales promelas days
Fathead minnow, 30
Pimephales promelas days
Fathead minnow, 4
Pimephales promelas mos
Fathead minnow, 4
Pimephales promelas mos
Channel catfish, 30
Ictalurus punctatus days
Channel catfish, 30
Ictalurus punctatus days
Channel catfish, 30
Ictalurus punctatus days
Channel catfish, 72
Ictalurus punctatus hrs
Channel catfish, 30
Ictalurus punctatus days
Channel catfish, 20
Ictalurus punctatus wks
Channel catfish, 2
Ictalurus punctatus wks
Channel catfish, 4
Ictalurus punctatus wks
LC50
LC50
Significant
mortality
Significant
mortality
Inhibition of
ATPase activity
Inhibition of
APTase activity
LC50
LC50
. LC50
Stimulated thyroid
activity
LC50
Weight loss and
liver hypertrophy
Increased trans-
aminase, lower
cortisol
Induced fish
hepatic microsomal
enzymes
4.7
A-1248
3.3
A-1260
23
A-1242
44 !
A-1016
0.31
A-1242
0.31
A-1254
75
A-1248
139
A-1254
433
A-1260
2.4 yg/g
A-1254
8.7
A-1242
20 wg/g
A-1242
8
A-1254
DeFoe, et al. In press
DeFoe, et al. In press ;
Herraanutz & Puglisi,-1976
Hermanutz & Puglisi, 1976
Cutkomp, et al.. 1972
Koch, et al. 1972
Mayer, et al. 1977
Mayer, et al. 1977
Mayer, et al. 1977
Mayer, et al. 1977
Mayer, et al. 1977
Hansen, et al. 1976
Camp, et al. 1974
1.000 Hill, et al. 1976
A-1254
-------
Table 7. (Continued)
Organism
Test
Duration Etfe
Result
(ug/i) Reterencfe
Flagfish,
Jordanella floridae
Mosquitofish,
Gambusia affinis
Mosquitofish.
Gambusia affinis
Guppy.
Poecilia formosa
Bluegill.
Lepomis macrochirus
Bluegill,
M Lepomis macrochirus
ro Bluegill.
to Lepomis macrochirus
Bluegill.
Lepomis macrochirus
Bluegill.
Lepomis macrochirus
Bluegill.
Lepomis macrochirus
Bluegill.
Lepomis macrochirus
Mink .
Mustela visbn
Mink.
Mustela vison
30
days
6
days
1.5
hr
1
day
5
days
•» w
^ •»
30
days
30
days
30
days
30
days
1
yr
Fin erosion
*
Bioconcentration
factor
Avoidance
Significant
mortality
LC50
Inhibit (150)
ATPase
Inhibition
of ATPase
LC50
LC50
LC50
LC50
Reduced
reproduction
Depressed
growth
37
A-1242
v 12,100
times
0.1
A-1254
200
A-1242
136
A-124S
0.6 yg/g
A-1242
30
A-1254
84
A-1242
78
A-1254
177
A-1254
400
A-1260
2 yg/g
10 pg/g
Hermanutz & Puglisi, 1976
Sanborn. 1974
Hansen, et al. 1974
Morgan, 1972
Mayer, et al. 1977
Desaiah, et al. . 1972
Yap. et al. 1971
Mayer, et al. 1977
Mayer, et al. 1977
Mayer, et al. 1977
Mayer, et al. 1977
Aulerich & Ringer, 1977
Aulerich, et al. 1973
Avoidance at 0.1 wg/1
-------
SALTWATER ORGANISMS
Introduction
Polychlorinated biphenyls (PCB's) were manufactured by the
direct chlorination of biphenyl; production in the United States
has now ceased. These mixtures were identified under the trade
name Aroclor and sold on the basis of percentage chlorine (e.g.,
21, 42,.54, and 60 percent). Since each component of the mixtures
differs in its physical, chemical, and biological properties, and
'^
since a possible 209 different chlorobiphenyls may be produced,
th,e .evaluation of the potential impact of the various mixtures on
the environment is complicated.
When an evaluation of the impact of PCB's on the environment
is performed, it is necessary to relate the data gathered in
laboratory, experiments with relatively pure mixtures to what
happens to the mixtures in nature. There is evidence that
percentages of chlorine change with time and location as the
mixtures are transported through the environment. For example,
the proportion of major peaks of Aroclor 1254 in shrimp and fish
captured from Escambia Bay, Florida differed from each other
(Nimmo et al., 1971). The major peaks in these organisms and in
organisms from laboratory studies (Hansen et al., 1971) also
differed from the standard used to calculate the amounts of the
chemical in tissues. Results of environmental monitoring by
Butler and Schutzmann (1978) showed that PCB's identified in
fishes, Pacific staghorn sculpin and English sole from the
Duwamish River, Washington State, during the period of fall 1972
B-23
-------
to spring 1976, changed from those resembling Aroclor 1254, to
those resembling Aroclor 1260, and later, Aroclor 1242.
Acute Toxicity
Acute toxicity tests of PCB mixtures to saltwater fishes have
not produced data that can be used to obtain 9.6-hour LC50 values
because concentrations tested were not sufficiently high (Table
12). Pinfish were not affected in 48 hours by 100 ug Aroclor
1254 per liter of water (Duke et al., 1970). Eighteen percent of
the pinfish died after 96 hours, compared to 2 percent of the
control fish, in water to which 100 ug Aroclor/1 was added (Hansen
et al., 1974a). Additional tests with saltwater fishes at
slightly higher concentrations might have given data necessary to
calculate 96-hour LC50 values. However, possible problems could
exist in validity of acute tests with PCB's because of their low
solubility in water (Schoor, 1975).
Available data suggest that saltwater invertebrate species
may be more acutely sensitive to PCB's than fishes (Table 8). The
adjusted LC50:or EC50 values for invertebrate species ranged from
2.8 to 12.0 ug/1; an unusually low variability in adjusted LC50
values. Because there was little difference in the toxicity of
different Aroclors, the geometric mean was calculated from all
adjusted LC50 values and divided by the species senstivity factor
of 49 to obtain a Final Invertebrate Acute Value of 0.14 ug/l«
The narrow range in adjusted LC50 values suggests that (1) a
species sensitivity factor of 49 is too great or (2) the more
B-24
-------
likely probability, based on freshwater invertebrate LC50 values
is that not enough species have been tested to establish the
variability, in sensitivity of saltwater invertebrate species.
Since there are too few data for PCB's and saltwater invertebrate
species to calculate a specific species sensitivity factor, the
guidelines value (49) is used. The Final Invertebrate Acute Value
is used as the Final Acute Value because, although LC50 values are
not available for fishes, they are not likely to produce a lower
acute value.
Chronic Toxicity
No life-cycle tests have been reported using saltwater
organism's. In an embryo-larval test (Table 9) with the sheepshead
minnow, fertilization was not affected by Aroclor 1254, but
significantly fewer embryos survived to hatching in a measured
concentration of 3.48 ug/1 (Schimmel et al., 1974). Survival of
fish during the two weeks following hatching was significantly
less in 0.16 ug/1 but not different from controls in 0.06 ug/1.
In a second study to determine the effect of PCB's in fish
eggs on survival, Hansen et al. (1973) exposed adult sheepshead
minnows for four weeks to Aroclor 1254 (Table 12). Adult fish
exposed to 5.6 ug/1 died but those in 1.1 ug/1 or lower apparently
were not affected. Embryos from adult fish were placed in
PCB-free flowing saltwater and observed for four weeks.
Fertilisation success was not affected by PCB's in eggs, but
survival of embryos and the resulting fry was reduced. Fry from
eggs containing 7.0 ug/g or more of PCB began dying a few hours
B-25
-------
after hatching. The concentration in eggs calculated to be lethal
to 50 percent of the fish was 6.1 ug/g. If PCS affects other
species similarly/ then other fish species with equally high
concentrations of Aroclor 1254 in their eggs may be endangered.
The effect of another PCB, Aroclor 1016, in water on fry,
juvenile or adult sheepshead minnows was determined in a four-week
exposure (Hansen et al., 1975). Survival of all three life-stages
was reduced in 15 u.g/1 but not in 5.5 ug/1 or less. Unlike
Aroclor 1254, as much as 77 u,g of Aroclor 1016/g of eggs
apparently did not affect survival of embryos and fry in water
free of this PCB.
Chronic exposure of fishes to Aroclor produced pathological
effects not observed in acute tests. Hansen et al. (1971)
reported signs of poisoning in pinfish exposed to Aroclor 1254,
such as fungus-like lesions on the body, hemorrhagic areas around
the mouth, ragged fins, etc. Signs of poisoning in sheepshead
minnows exposed to Aroclor 1254 included lethargy, reduced
feeding and fin rot (Hansen et al., 1973; Schimmel et al., 1974).
Spot exposed to Aroclor 1254 for two weeks or longer showed
fatty changes in their livers (Nimmo et al., 1975). In inter-
mediate stages of liver pathogenesis in fish exposed to Aroclor
1254, there were extreme fatty changes characterized by the
presence of large vacuoles within hepatocytes and disorientation
of liver cord distribution. In advanced stages of pathogenesis in
a moribund fish, there were intracellular PAS-positive bodies
(ceroid), congestion of blood sinuses, and severe vacuolation.
B-26
-------
Life-cycle tests with the fathead minnow and Aroclors 1242,
-
-------
alterations:'in the vesicular connective tissue (parenchyma) around
the digestive diverticular of the hepatopancreas.
Aroclor 1254 was toxic to the saltwater amphipod, Gammarus
•
oceanicus,. at a nominal concentration of 10.0 ug/1 (Wildish,
*. •.
1970) . Molting animals were particularly vulnerable to the PCS.
Necrotic branchia were found in some animals exposed for about 6
days to a nominal concentration of 1.0 ug/1.
Arpclor 1254 affected the species composition of communities
of estuarine animals that developed from planktonic larvae in salt'
water that flowed for four months through small aquaria (Table 12;.
Hansen, 1974) . The number of arthropods decreased while the
number of :;ehordates increased in aquaria receiving 0.6 ug/1 of the
PCS. Numbers of phyla, species and individuals were decreased by
this PCB, but there was no apparent effect on the abundance of
annelids, brachiopods, coelenterates, echinoderms, or nemerteans. '
This study showed that a PCB can have marked effects on community
structure at concentrations not much different from those that
produced chronic effects on single species.
No Final Invertebrate Chronic Value can be obtained because
no appropriate chronic"tests on saltwater invertebrate species
were found in the literature. However, extended exposures of
saltwater species and life-cycle tests with freshwater inverte*-
brate species.(Table 4) demonstrate that acute tests underestimate
* . * •*
the chronic toxicity of PCB's (Table 12). Therefore, knowledge of
the chronic effects of this PCB are .critical to the generation of
a criterion.
B-28
-------
Plant Effects
Information concerning the sensitivity of plants is
restricted to unicellular algae (Table 10). Fisher and Wurster
.(1973) found that the growth of the diatom, Rhizosolenia setigera,
was reduced in a medium to whch 0.1 u.g/1 Aroclor 1254 was added.
Likewise, Fisher et al. (1974) demonstrated that 0.1 ug Aroclor
1254 added per liter of water changed the species ratio of the
alga,fDunaliella tertiolecta, and the diatom, Thalassiosira
pseudonana. Fisher et al. (1974) also showed a decrease in
species diversity and species ratio change in natural phytoplank-
ton communities at 0.1 ug/1 Aroclor 1254. In summary, some data
suggest that unicellular plants are affected by concentrations of
PCB's similar to concentrations that are chronically toxic to
animals. Unfortunately no data using measured concentrations were
presented and it is difficult to interpret the ecological
significance of these studies.
Bioconcentration
The bioconcentration factors (BCF's) of PCB's in saltwater
species in laboratory tests are shown in Table 11. The diatom,
Cylindrotheca closterium, had a bioconcentration factor of 1,000
(Keil et al., 1971); Eastern oyster, up to 168,000 times (Lowe et
al., 1972); grass shrimp, Palaemonetes pugio, 42,000 times (Nimmo
• ' ' *-
et al., 1974), and in the three fishes listed, Leiostomus
xanthurus, Cyprinodon variegatus, and Lagodon rhomboides, as high
B-29
-------
as 44,000 times (Hansen et al., 1971; 1974a, and 1974b).
Bioconcentration factors for PCB's in six of seven species of
freshwater fishes in laboratory tests were generally similar,
ranging from 5,000 to 60,000.
Bioconcentration factors calculated from data from Escambia
Bay, Florida were greater than 230,000 for blue crabs (Nimmo et
al., 1975), and greater than 100,000 for oysters; :and 670,000 for
speckled trout (Duke et al., 1970). These data, and field data on
freshwater fishes, suggest that bioconcentration factors from
laboratory studies underestimate bioconcentration potentials of
PCB's in the environment (Hansen, 1975).
The bioaccumulation of PCB's into aquatic organisms from
PCB's in food and in water and the effects of PCB's on mammals
that feed on fish and shellfish are important. The lowest maximum
permissible tissue concentration (0.64 ug/1) is based on the
effect of dietary PCB's on mink (Platonow and Karstad, 1973).
Using the geometric mean fish bioconcentration factor (27,000) a
Residue Limited Toxicant Concentration of 0.024 ug/1 is obtained.
Effects on mink were seen at a dietary PCB concentration of 0.64
ug/g and a "no-effect" dose was not determined. A criterion
calculated from these data may not be protective because the
dietary concentration was not protective and the BCF based on
laboratory studies may underestimate BCF's in saltwater animals
since field-observed bioconcentration factors were higher but
could not be used in the calculations. When field data were used
for freshwater fish, a much higher BCF (427,000) was derived.
B-30
-------
Miscellaneous
, No other data exist that suggest any more sensitive effects
(Table 12).
B-31
-------
CRITERION FORMULATION
Saltwater Aquatic Life
Summary of Available Data .c -^r_?:.I*z,-=•
The concentrations below have be'en rounded to twd signif-
icant figures. -
Final Fish Acute Value = not available
Final Invertebrate Acute Value = 6.20'jag/1
Final Acute Value = 0.20 jig/I
Final Fish Chronic Value = 0.049 ug/1
Final Invertebrate Chronic Value = not available
Final Plant Value =0.1 ug/1
Residue Limited Toxicant Concentration = 0.024 jig/I
Final Chronic Value = 0.024 ug/1
0;44 x Final Acute Value = 0.087 jig/1
The maximum concentration of polychlorinated biphenyls
is the Final Acute Value of 0.20 ug/1 and the 24-hour average
concentration is the Final Chronic Value of 0.024 pg/1.
No important adverse effects on saltwater aquatic organisms
have been reported to be caused by concentrations lower
than the 24-hour average concentration.
CRITERION: For polychlorinated biphenyls the criterion
to protect saltwater aquatic life as derived using the Guide-
lines is 0.024 ug/1 as a 24-hour average and the concentration
should not exceed 0.20 pg/1 at any time.
B-32
-------
Table 8. Marine invertebrate acute values for polychloririated biphenyls
Adjusted
00
1
u>
• Or nanism
Eastern oyster,
Crassostrea virginica
Eastern oyster,
Crassostrea virginica
Eastern oyster.
Crassostrea virginica
Eastern oyster.
Crassostrea virginica
Brown shrimp,
Penaeus aztecus
Grass shrimp,
Palaemonetes pugio
Pink shrimp.
Penaeus duorarum
Pink shrimp.
Penaeus duorarum
Bicxissay
Method*
FT
FT
FT
•
FT
FT
FT
FT
FT
Test
cone . **
^MIMWri^V^B
U
u
u
u
u
u
u
u
Chemical
Description
A- 1016
A- 1248
A- 1254
A-1260
A- 1016
A- 1016
A- 1248
A-1254
•
Time
(hrs)
96
24
24
24
96
96
48
48
LCbO
(qq/1)
10.2***
17. ***
14.0***
60. ***
10.5
12.5
32. ***
32.0***
LCbu
(uq/1)
7.8
3.4
2.8
12.0
8.1
9.6
10.5
10.5
Reference
Hansen, et
al. 1974a
Lowe, undated
Lowe, undated
Lowe, undated
Hansen, et
al. 1974a
Hansen, et
al. 1974a
Lowe, undated
Lowe, undated
* FT =» flow-through
** U = unmeasured
***EC50: Decreased growth of oysters; loss of equilibrium or death of shrimp.
Geometric mean of adjusted values =9.67 ug/1 ^T =0.20 ug/1
-------
Table 9. Marine fish chronic values for polychlorinated biphenyls
''Organism'
Sheepshead minnow,
Cyprinodon variegatus
Sheepshead minnow,
Cyprinodon variegatus
Test* ;
E-L
Limits
(uq/Jl>
Chronic
Value
3.4-15.0** 3.6
0,06-0.16 0.049
Reference
Hansen. et al. 1975
Schimmel, et al. 1974
u»
* E-L = embryo- larval test
** Aroclor 1016
***Aroclor 1254
Geometric mean = 0.42 pg/l
Lowest chronic value «• 0.049 pg/1
0.060 ng/1
-------
Table 10. Marine plant effects for polychlorinated biphenyls
CO
I
(jj
ui
Organism
Diatom,
Effect
No growth in
Concentration
(aq/il
0.1*
Rhizosolenia setigera 48 hr. Reduced
growth thereafter
Diatom.
Thalassiosira
pseudonana
Diatom,
Skeletonema costatum
Diatoms,
Thalassiosira
§seudonana7
keletonema
Reduced growth 25-100*
costatum
Diatom,
Cylindrotheca
closCerium '
Phytoplankton
populations
Green alga.
Reduced growth
Reduced growth
and carbon
fixation in
48 hr
Reduced growth
Toxicity in
24 hr
Species ratio
Dunaliella tertiolecta/ change
Diatom
Thalassiosira pseudonana
(mixed culture)
Green alga, Species ratio
Dunaliella tertiolecta/"change
Diatom
Thalassiosira pseudonana
(mixed culture;
Natural phytoplankton Decrease in
10*
10*
100**
15*. 6.5**
1*
0.1*
0.1*
community
Alga.
species diversity,
species ratio
change
Reduction in rate 100*
Dunaliella tertiolecta of carbon fixation
Reference
Fisher & Uurster. 1973
Mosser, et al. 1972a
Mosser, et al. 1972a
Fisher, 1975
Keil, et al. 1971
Moore & Harriss, 1972
Mosser, et al. 1972b
Fisher, et al. 1974
Fisher, et al. 1974
Luard, 1973
* Aroclor 1254
** Aroclor 1242
Lowest plant value v 0.1 pg/1
-------
Tatle 11. Marine residues for polychlorinaced biphenyls
00
I
Organism
Diatom,
Cylindrotheca closterium
Eastern oyster,
Crassoscrea virginica
Eastern oyscer,
Crassostrea virginica
Eastern oyster,
Crassostrea virginica
Grass shrimp,
Palaemonetes pugio
Blue crab,
Callinectes sapidus
Spot,
Leiostomus xanthurus
Sheepshead minnow,
Cyprinodon variegatus
Pinfish,
Lagodon rhotnboides
Speckled trout,
Cynoscion nebrelosus
Fishes
Invertebrates
Bioconcentration factor
1,000*
13,000**
101,000***
>100,000*****
27,000***
>230,000*****
37,000***
30,000***
17,000*
>670,000*****
>133.000****
>27,000****
Time
(days)
14
84
245
neference
Keil, et al. 1971
Parrish. et al. 1974
Lowe, et al. 1972
Field data Duke, et al. 1970
16
Field data
28
28
21-28
Field data
Field data
Field data
Nimmo, et al. 1974
Nimmo, et al. 1975
Hansen, et al. 1971
Hansen, et al. 1973
Hansen, et al. 1974a
Duke, et al. 1970
Nimmo, et al. 1975
Nimmo, et al. 1975
* Aroclor 1242
** Aroclor 1016
*** Aroclor 1254
**** Averages from field data from Escambia Bay, Fl., based on 27 water samples, 101 invertebrate
samples, and 17 fish samples expressed, as Aroclor 1254.
***** Greatest bioconcentration factor of Aroclor 1254 in moHusks, crustaceans, or fishes from
Escambia Bay, Florida.
-------
Table 11 (continued)
Organism
Bioconcentratlon Factor
Time
(days)
i
-------
Table 12. Other marine data for polychlorinated biphenyls
03
I
LJ
00
Organism
Ciliate protozoans,
Tfctrahymena pyriformis
Ciliate protozoan,
Tetrahymena pyriformis
Ciliate protozoan,
Tetrahymena pyriformis
Ciliate protozoan,
Tetrahymena pyriformis
Eastern oysters,
Crassostrea virginica
Eastern oysters,
Crassostrea virginica
Horseshoe crab,
LimuLus polyphemas
Amphipod,
Gammarus oceanicus
Grass shrimp,
Palaemonetes pugio
Pink shrimp,
Penaeus duorarum
Pink shrimp,
Penaeus duorarum
Pink shrimp,
Penaeus ouorarum
Fiddler crab,
Uca pugilator
Communities of
organisms
Spot,
Leiostomus xanthurus
Test
Duration Ettect
7 days Bioconcentration
factor = 60*
96 hrs Reduced growth
96 hrs Reduced growth
96 hrs Reduced growth
2 days Bioconcentration
factor = 8,100*
24 wks Reduced growth
96 days Bioconcentration
factor = 1.298****
30 days Mortality
Result
tug/it Reference
1 hr
Avoidance
15 days 51% mortality
15 days LC50
2 days Bioconcentration
factor = 140*
38 days Inhibited molting*****
4 mos Affected composition
12 days 50% mortality
Cooley, et al. 1972
1000.** Cooley, et al. 1973
1.0* Cooley, et al. 1972
1000.*** Cooley, et al., 1973
Duke, et al. 1970
5.0* Lowe, et al. 1972
Neff & Giam, 1977
>10.0* Wildish. 1970
<100.0*
10.0* Hansen, et al. 1974b
0.94* Nimmo, et al. 1971
1.0* Nimmo & Bahner, 1976
Duke, et al. 1970
8.0* Fingerman & Fingerman,
1977
0.6* Hansen. 1974
5.0* Hansen, et al. 1971
-------
Organism
Pinfish.
Lagodon rhomboldes
Pinfish.
Lagodon rhomboides
Pinfish,
Lagodon rhomboides
Pinfish.
Lagodon rhomboides
Table 12. (Continued)
Test
Duration Etfect
• 1 hr Avoidance
18 days 50% mortality
2 days Bioconcentration
factor = 980*
42 days 507. mortality
Result
Jua/il • Reference
10.0* Hansen. et al. I974b
5.0* Hansen. et al. 1971
Duke, et al. 1970
21.0**** Hansen, et al. 1974a
Sheepshead minnow. 28 days Affected reproduction*** 0.14* Hansen, et al. 1973
Cyprinodon variegatus
00
I
W
vo
* Aroclor 1254
** Aroclor 1248
*** Aroclor 1260
**** Aroclor 1016
*****Aroclor 1242 '
******significantly affected hatching of eggs or the survival of fry from exposed adults.
-------
POLYCHLORINATED BIPHENYLS
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-------
.PCBS
Mammalian Toxicology and Human Health Effects
SUMMARY .,
Polychlorinated biphenyls (PCBs) have been used com-
mercjLadly since 1929 as dielectric and heat exchange fluids
and in a variety of other applications. They have become
widely disseminated in the environment worldwide. Like
many organochlorine pesticides, they are highly persistent
and accumulate in food webs. Human exposure to PCBs has
resulted largely from the consumption of contaminated food
but also from inhalation and skin absorption in work environ-
ments. PCBs accumulate in the fatty tissues and skin of
man and other mammals. Metabolism occurs by hydroxylation
and dihydrodiol formation with arene oxides as probable
intermediates. The rate of metabolism and excretion slows
dramatically as the chlorination of the biphenyl nucleus
increases. Arrangement of chlorines which eliminate adjacent
unsubstituted carbons greatly increase resistance to metabolism.
PCBs have caused profound toxic effects in man and animals,
particularly if repeated exposures occur. The skin and
liver are major sites of pathology with the gastrointestinal
tract and nervous systems also being targets. Polychlorodi-
benzofurans which contaminate commercial PCB mixtures may
contribute significantly to their toxicity. Several studies
in rodents suggest strongly that some PCBs are carcinogenic
C-l
-------
and that they can enhance the carcinogenic!ty of other chemi-
cals. A linear model for risk assessment has been used
to estimate maximum safe levels in water and fish which
will establish a level of risk for the human population
from cancer. A maximum level of PCBs in water projected
to result in no more than 1 cancer in 10 individuals with
lifetime exposure of 0.26 ng/liter is suggested by the analysis.
C-2
-------
EXPOSURE
Introduction
PCB's have t>ecome widespread in the environment since
the introduction of their commercial use in 1929 (Peakall,
1975). The magnitude of the dispersal of these chemicals
is revealed by their detection in the tissues of plants
and animals in all parts of the world. PCB residues have
been observed in wildlife in Sweden, North America, Great
Britain, the Netherlands, and even the Arctic (Risebrough
and de Lappe, 1972). Because PCBs are not naturally occurring
substances, their dissemination is entirely the result of
human activity. Their entry into the environment has occurred
by vaporization into the atmosphere, and by spilling or
dumping into water or onto land. It has been estimated
that of the 1970 sales of PCBs in North America only 20
percent represented a net increase in the total amount in
service. Estimated sources of loss for that year were 1
- 2 x 10 tons for evaporation; 4 - 5 x 10 tons for leaks
and disposal of fluids; and 22 x 10 tons from disposal by
incineration and burial (Nisbet and Sarofim, 1972). The
cumulative input to the environment between 1930 and 1970
4 4
was estimated to be 3 x 10 tons to air, 6 x 10 tons to
fresh and coastal waters, and 3 x 10 tons to dumps and
landfills. In that time, up to 1/3 of the PCBs released
to air and 1/2 of that released to water were probably degrad-
ed. Degradation in landfills is more difficult to estimate
(Nisbet and Sarofim, 1972). PCBs have been found repeatedly
to be widespread in analyses of human tissues. For example,
detectable levels of PCBs have been reported in adipose
C-3
-------
tissue samples of up to 91 percent of individuals sampled
in a survey of the United States population (Kutz and Strassraan,
1976; see Table 12). This finding suggests that environmental
contamination may be a significant source of human exposure.
Likely routes of exposure for the general population are
water and particularly food while inhalation and dermal
icontact are likely to be more significant routes in occupa-
tional exposure.
C-4
-------
Ingestion from Water
The solubility of PCBs in water is very low, decreasing
as the percent chlorination is raised. Solubilities of
Aroclors in water at 20° vary from 200 >ug/l for 1242 to
about 25 ;ug/l for 1260 (Nisbet and Sarofim, 1972). The
major factors in the dynamics of PCB distribution in water
are its low solubility, high specific gravity, and its high
affinity for solids. Most PCBs discharged into water are
found in bottom sediments near the site of discharge (Nisbet
and Sarofim, 1972). Evaluation of PCB levels in surface
waters and bottom sediments of the major drainage basins
of the United States was conducted between 1971 and 1974
(Dennis, 1976). The data were derived from the U.S. Geological
Survey study of 1971-72 (Crump-Weisner, et al. 1974) and
from additional data collected by the USGS between 1972
and 1975 (PCB data base 1972-75). It is summarized in detail
in the Criteria Document for PCBs (U.S. EPA, 1976a). The
highest concentrations in both water and sediment were found
in the basins east of the Mississippi River. The highest
levels were found in 1971 in the lower Mississippi basin
with a median concentration for the region of 3 ug/1 and
positive detections at 100 percent of stations tested.
Over the time period of the study the concentrations and
incidences of PCBs detected in all basins have decreased
substantially. By 1974 the median level in the lower Missis-
sippi basin had dropped to 0.1 jug/1 and the incidence of
detection to 2.6 percent of stations tested. The levels
in sediments however, have persisted at much higher levels
c-5
-------
over this period of time. In 1971 median sediment levels
for the Mississippi basin were 30 jug/kg and the incidence
of detection 100 percent. By 1974 the incidence had dropped
to 9.9 percent and the median level was 10.5 jug/kg.
Although PCBs are widespread in aquatic environments
(Peakall, 1975) their low solubility generally prevents
them from reaching high concentrations in drinking water
.supplies. The persistence of PCBs and their accumulation
in sediments increase the significance of water as a source
of human exposure by providing a reservoir of material which
can continue to contaminate water long after the addition
of PCBs has ceased. In combination with these factors,
the lipophilicity of PCBs results in their continued intro-
duction to, and accumulation in, the food chain. As a con-
sequence, fish and other foods obtained from aquatic environ-
ments may become important sources of exposure even if PCB
levels in the water are low.
The ability of PCBs discharged from a manufacturing
facility to contaminate a drinking water system has recently
been highlighted. Billings, et al. (1978) determined the
levels of PCBs in the Easley-Central Water District, Pickens
County, South Carolina. They observed that PCBs discharged
by a capacitor manufacturing facility 12 km upstream from
the water district's treatment plant were entering the water
system. Finished potable water supplies were contaminated
to levels as high as 818 ng/1.
C-6
-------
Ingestion from Foods
Contamination of food with PCBs occurs primarily by
three mechanisms. The first is contamination of human food
as a consequence of accumulation in the food chain. The
contamination of freshwater fish as a consequence of the
contamination of the aquatic environment is a particularly
significant route of PCB entry into the human diet which
will be discussed in more detail below. The second mechanism
occurs by the direct contamination of feeds or foodstuffs
with PCBs. This may occur as a result of accidental spills
or equipment malfunctions as was the case in the episode
of rice oil contamination in Japan which led to the outbreak
of Yusho or rice oil disease in 1968 (Kuratsune, et al.
1976). In this instance leaks in a heat exchanger used
to process rice bran oil resulted in the contamination of
the oil by the exchanger fluid (Kanechlor 400) . Discovery
of the contamination was made only after numerous cases
of chlorinated hydrocarbon intoxication in Fukuoka prefecture,
Japan. The oil was found to contain 2,000 to 3,000 ppm
Kanechlor 400 which was contaminated with polychlorodibenzo-
furans (1.6 to 5 ppm). Average consumption of PCBs among
affected individuals was estimated to be 2 g (Kuratsune,
et al. 1972). By 1975 the total number of known individuals
affected was 1,291. Elevated PCB levels in fat were still
observed four years after the exposure, and dermatological
symptoms were found in up to 89 percent of a group of 72
patients examined in 1973 or 1974. Another example of acci-
dental PCBs contamination in animal feed occurred as a result
C-7
-------
of the use of PCBs in silo coatings (Willett and Hess, 1975),
The third significant source of PCBs in foodstuffs was food
packaging made from recycled paper containing PCBs (Jelinek
and Corneliussen, 1976).
A special case of human exposure via food which must
be considered is human breast milk. Adverse effects have
been observed in breast fed infants of women with Yusho
(Kuratsune, et al. 1976) and in infant Rhesus monkeys ingest-
ing breast milk containing 7 to 16 ppm PCBs (fat basis)
(Allen, 1975; Allen and Barsotti, 1976). Preliminary survey
data indicate average PCB levels in human breast milk of
1.8 ppm (fat basis) (42 FR 17487) and a study of PCB exposed
nursing mothers in Germany indicated average PCB levels
of 3.5 ppm (Tombergs, 1972). The proximity of these values
to the toxic levels in infant monkeys (7 to 16 ppm) suggests
that human breast milk must be considered a significant
source of PCB exposure.
The extent of contamination of the U.S. food supply
has been the subject of Food and Drug Administration (FDA)
and Department of Agriculture (USDA) monitoring programs
since 1969. Results of these studies have been summarized
by Jelinek and Cornelliussen (1976). The initial analysis
of 15,000 food samples between 1969 and 1971 is summarized
in Table 1. The results of monitoring programs in fiscal
years 1973, 1974, and 1975 are summarized in Table 2. Over
the monitored period the incidence and levels of PCBs have
dropped in all food classes. By 1975 the only significant
food sources were fish, meat, and dairy products. Fish were
C-8
-------
TABLE 1
Summary of PCB's in Food
Nov., 1969 - June, 1971a
Food
commodity
Finf ish
Oysters
Fish byproducts
Cheese
Milk
Eggs
Potato
byproducts
Miscellaneous
Positive
findings
317
12
6
44
60
17
12
11
Avg. of
positives
(ppm)
2.1
Trace
1.8
0.3b
2.5b
Trace
1.1
1.9
Max. level
(ppm)
35.3
Trace
5.0
1.0b
22. 8b
0.5
4.2
6.5
Approximately 15,000 samples examined.
Fat basis.
°Detection limits: fish 0.5 ppm, other foods 0.05 ppm (P.E. Corneliusser
personal communication).
From: Jelinek and Corneliussen (1976).
C-9
-------
TABLE 2
Summary of PCBs in Foods. FY73, 74 and 75
Food
Commodity
Fish
Milk
Eggs
Cheese
Feed components
Animal feeds
Processed fruits
Infant & jr. foods
Meats & poultry
(USDA)
FY '
Percent
positive
60.4
2.2
1.1
0.9
12.7
7.2
4.5
1.1
Percent
positive
1.9
73
Max.
(ppm)
123.0
1.6
Trace
0.5
9.0
199.5
19.2
Trace
Percent
above
5 ppma
0.19
FY
Percent
positive
44.0
2.6
4.2
2.6
0.0
0.0
0.0
0.0
Percent
positive
1.2
'74
Max.
(ppm)
16.8
2.3
11.0
2.8
N.D.
N.D.
N.D.
N.D.
Percent
above
5 ppma
0.07
FY
Percent
positive
17.8
0.7
0.0
0.0
0.3
0.0
0.0
0.0
Percent
positive
0.3
'75
Max.
(Ppm)
9.0
1.9
N.D.
N.D.
0.9
N.O.
N.D*
N.D.
PercVr
above .
5 ppm'
0.06
aMilk, cheese, meats and poultry reported as ppm, fat basis.
Detection limits: fish 0.5 ppm, other foods 0.05 ppm. (P.E. Corneliussen,
personal communication).
From: Jelinek and Corneliussen (1976).
C-10
-------
by far the most significant source. The findings for the
1969-71 period led to the establishment of regulations for
PCB levels in food (38 FR 18096). The temporary tolerances
established at that time and new tolerances recommended
in 1977 (42 FR 17487) are given in Table 3. The enforcement
of those tolerances and restriction of PCB use in open systems
after 1970 probably account for the general decline of PCB
levels in foodstuffs.
Comprehensive fish surveys conducted by the FDA in
fiscal years 1973 and 1974 indicated a drop in the incidence
of PCB detection in fish from less than 30 percent in 1973
to less than 20 percent in 1974. In 1973 three percent
contained over 1 ppm and 0.5 percent contained over 5 ppm
PCBs. The data from all FDA studies in the fiscal years
1973, .1974, and 1975 are summarized in Figure 1. While
the incidence of PCBs in fish dropped over the period the
fraction of positive fish containing over 5 ppm PCBs increased
from less than five percent to over ten percent. The samples
containing more than 5 ppm were from the Great Lakes. Because
the study involved different sources and objectives from
year to year no conclusion as to whether a significant trend
existed was drawn. It should be noted that these surveys
were conducted with fish in commerce and provide no information
about sport fish per se. The studies indicated that signi-
ficant levels of PCBs generally do not occur in salt water
fish.
C-ll
-------
60
50
10
30
S 20-
10
n
i
yj
g\
FISCAL YEAR '/J
FISCAL YEAR '7<4
FISCAL YEAR '75 »
Figure 1: PCBs in fish FY73, 74, 75 a level of detection:
0.5 ppm.
Prom: Jelinek and Corneliussen (1976).
C-12
-------
TABLE 3
FDA Regulations for PCBs
I.
Commod i ty
Milk (fat basis)
Dairy products (fat basis)
Poultry (fat basis)
Eggs
Finished animal feed
Animal feed components
Fish (edible portion)
Infant and junior foods
Temporary tolerances
PCB cone.
(ppm)
2.5
2.5
5.0
0.5
0.2
2.0
5.0
0.2
Proposed Guidelines
1977
1.5
1.5
3.0
0.3
0.2
2.0
2.0
pending
Paper food-packaging material 10. Oa
without PCB-impermeable barrier
II. Use prohibited in food,
food packaging plants
feed,
Administrative guideline, pending hearing.
From: Jelinek and Corneliussen (1976)
42 FR 17487
C-13
-------
The impact of sport fish consumption was examined in a study
of a group of sports fishermen who consumed an average of
24 to 25 pounds of fish annually (highest individual exposure
180 Ibs/year over a two-year period). PCS residues in cooked
fish ranged from 0.35 - 5.38 ppm. Plasma PCB levels ranged
from a high of 0.366 ppm in the exposed group to control
levels 0.007 ppm (less than six Ibs consumed per year) (42
FR 17487) -.
A bioconcentration factor (BCF) relates the concentration
of a chemical in water to the concentration in aquatic organisms,
but BCF's are not available for the edible portions of all
four major groups of aquatic organisms consumed in the United
States. Since data indicate that the BCF for lipid-soluble
compounds is proportional to percent lipids, BCF's can be
adjusted to edible portions using data on percent lipids
and the amounts of various species consumed by Americans.
A recent survey on fish and shellfish consumption in the
United States (Cordle, et al. 1978) found that the per capita
consumption is 18.7 g/day. From the data on the nineteen
major species identified in the survey and data on the fat
content of the edible portion of these species (Sidwell,
et al. 1974) , the relative consumption of the four major
groups and the weighted average percent lipids for each
group can be calculated:
C-14
-------
Group
Freshwater fishes
Saltwater fishes
Saltwater molluscs
Saltwater decapods
Consumption
(Percent)
12
61
9
18
Weighted Average
Percent Lipids
4.8
2.3
1.2
1.2
Using the percentages for consumption and lipids for each
of these groups, the weighted average percent lipids is
2.3 for consumed fish and shellfish.
Measured bioconcentration factors were obtained in
tests lasting over 200 days with three species using Aroclor
1242, 1248, 1254, and 1260:
Organism
Brook trout
Salvelinus fontinalis
fillets, A-1254
whole body, A-1254
Average
BCF
2,800
12,000
Percent
lipids
0.65
2.8
Adjusted
BCF
9,900
9,900
Reference
Snarski &
Puglisis,
1976
Fathead minnow
Pimephales promelas
females, A-1248
males, A-1248
males, A-1260
females, A-1260
Fathead minnow
Pimephales promelas
males, A-1242
females, A-1242
males, A-1254
females, A-1254
240,000
120,000
270,000
540,000
123,000
75,000
181,000
283,000
10.4
4.2
3.3
9.7
3.8
10.0
3.8
10.0
53,000
66,000
188,000
128,000
74,000
17,000
109,000
49,000
DeFoe, et al.
1978
Nebeker, et
al. 1974
C-15
-------
Eastern oyster Lowe, et al,
Crassostrea yirginica 1972
A-1254 93,000 1.5 142,000
Only tests lasting over 200 days were used since long exposures
are necessary to reach steady-state. The percent lipids
for mature fathead minnows were obtained from DeFoe, et
al. (Personal communication). The percent lipids for oysters
was obtained from Sidwell, et al. 1974. Each of these average
measured BCF's was adjusted from the percent lipids of the
test species to the 2.3 percent lipids that is the weighted
average for consumed fish and shellfish. The geometric
mean was obtained for each species, and then for all species.
Thus, the weighted average bioconcentration factor for PCB's
and the edible portion of all aquatic organisms consumed
by Americans is calculated to be 46,000.
Higher BCF values apparently can be achieved in field
exposures (Haile, et al. 1975; Norstrom, et al. 1976; Duke,
et al. 1970; Nimmo et al. 1975; Veith, 1975; Veith, et al.
1977) , but those values cannot be considered quantitative
because the exposure of the organism cannot be adequately
documented and integrated over a long enough period of time.
In order to estimate human dietary PCB intake the FDA
conducts a continuing survey of the total diet. Composites
of 12 different food categories are analyzed for PCB content.
Table 4 summarizes the results of the survey from 1971 through
the first half of 1975. While contamination was observed
in most categories in 1972 the number of positive categories
had dropped by 1973. In 1974 and 1975 only meat, fish,
and poultry were observed to contain PCBs and fish was almost
C-16
-------
TABLE 4
Percent of Composites Containing PCBs,
From the FDA Total Diet Studies
Fiscal Year
1971
1972
1973
1974
1975 (1st half)
6
10
•
CO
•o
£ 0
CO >,
••-1 Vl
fa 4J
iH
•• 3
JJ 0
(U <1)
J >
(0
0)
__J
^^
X!
(0
-U
4J (U
O CT
O (U
« >
C CO
0) -U
T3 -iH
>-i 3
(Q W
O fa
C
(0 —i
4J C
O
x:
m to
4J
to o
Vi C
ro =1
cr>Ti
3-D
co <
17
3
From: Jelinek and Corneliussen. 1976.
C-17
-------
always the contributor of positive results in that category
(Jelinek and Corneliussen, 1976). Most of the contamination
noted in the other categories in earlier years was thought
to result from exposure during processing or packaging be-
cause the raw foods were rarely found to contain PCBs.
Total daily intake, calculated from the composite figures
for a young adult male over the period 1971-1975, is sum-
marized in Table 5. Total daily intake dropped by almost
50 percent over the period but intake in the meat-fish-poultry
category changed very little. By 1974 almost all of the
dietary intake resulted from the ingestion of PCB-contami-
nated fish. The measures taken in the early 1970's to limit
the release of PCBs into the environment and to remove them
from food processing environments effectively reduced direct
contamination of foodstuffs to a minimum level. The per-
sistence of PCBs in aquatic environments and in fish has
maintained a residual dietary exposure level in the diet.
Further reduction of PCB levels in the diet will require
that entry of PCBs into waterways be more tightly controlled
and that monitoring of fish and other foods for PCB contami-
nation be continued (Jelinek and Corneliussen, 1976). The
recently recommended reduction of allowable PCB levels in
fish to 2.0 ppm may further reduce dietary intake (42 PR
17487).
Two special situations should be mentioned which must
be avoided to prevent unnecessary PCB ingestion. First,
accidental contamination of foodstuffs or feeds with PCBs
must be avoided. Although PCB manufacture is now stopping
c-18
-------
TABLE 5
Estimates of daily PCB intakes
(Total Diet; Study - Teenage Male)
Average daily intake of PCB' sa
Fiscal
year
1971
1972
1973
1974
1975. (1st half)
Total diet
(jug/day)
15.0
12.6
13.1
8.8
8.7
Meat-fish- poultry
food class (jug/day)
9.5
9.1
8.7
8.8
8.7
aLower limit of quantitative reporting =
0.05 ppm with analytical method employed.
From: Jelinek and Corneliussen (1976)
C-19
-------
and distribution will cease in the near future, many PCB
containing products remain in service. Failure to exercise
care in the maintenance and disposal of these units could
result in the contamination of food or water. The tragic
results of the episode of rice oil contamination in Japan
(Kuratsune, 1972) provides ample evidence of the need for
care and continued surveillance of foods. Second, although
occupational exposure to PCBs will decline over the next
several years, the possibility of food contamination as
a consequence of transfer from workers tools' or clothing
must be considered as a possible route of dietary exposure.
Inhalation
PCBs can enter the atmosphere by vaporization and may
be found in either gaseous form or adsorbed to airborne
particulates. Prior to the restriction of PCB use, substan-
tial losses to the atmosphere resulted from evaporation
of plasticizers and from improper incineration (Nisbet and
Sarofim, 1972) . In 1972 terrestrial input from fallout
was estimated to be 1000 to 2000 tons/year. Annual emission
rates were estimated at 1500 to 2500 tons. (Nisbet and Sarofim,
1972). In 1975 a study of PCB content in air in suburban
areas in Florida and Colorado indicated that average atmos-
pheric levels were approximately 100 ng/m (Kutz and Yang,
1976) . Rates of fallout along the southern California Coast
2
were estimated to average 1800 kg/year over a 50,000 km
area (Young, et al. 1976). The distribution of PCBs in
air is non-uniform, being more highly concentrated in urban
areas. The aerial fallout survey in southern California
C-20
-------
indicated that sectors in the urban areas around Los Angeles
had fallout rates of up to 180 kg/yr while less industrialized
sectors had rates as low as 30 kg/yr. A study of PCS levels
in soil samples showed that they were rarely detectable
in agricultural soils but were found in 63 percent of urban
samples from 19 cities (Carey and Gowan, 1976). General
human exposure to inhaled PCBs probably varies with the
local conditions. Relative to the 9 /ig/day intake estimated
from the diet (Jelenek and Corneliussen, 1976) non-occupa-
tional exposures by inhalation are probably small.
While inhalation of PCBs is not and most likely will
not be a major route of general human exposure, it is a
highly significant route of occupational exposure. Early
in its commercial use an association was observed between
occupational exposure to PCB vapors and chloracne (Jones
and Alden, 1936; Schwartz, 1936). The benefits of control-
ling leaks from closed systems into work environments were
noted by Meigs, et al. (1954).
A study of occupational exposure in Japan found PCB
vapors at levels between 13 and 540 jug/m and airborne par-
ticulates between 4 and 650 jug/m in a survey of six indus-
trial plants. An additional finding of 6,270 jjg/m PCB
particulates was associated with a spill. Blood PCB levels
of 99 exposed workers averaged 370 ppb as compared to levels
in 32 controls of 20 ppb (Hasegawa, et al. 1972). Ouw,
et al. (1976) observed Aroclor 1242 levels between 2.22
and 0.32 mg/m in different areas of an electrical equipment
I
manufacturing facility in Australia. Blood Aroclor levels
c-21
-------
were analyzed by gas chromatography and fractions with several
retention times standardized against Aldrin were detected
in exposed workers. Workers in an impregnation room where
inhalation was a major mode of exposure had higher levels
of PCBs than did workers in another area where exposure
was primarily dermal. A series of 30 control individuals
were not found to have detectable PCB levels. The limit
of detection in this study was not reported; however, Finklea,
et al. (1972) reports American control population blood
levels of 0.3 to 3 ppb.
It is difficult to differentiate between industrial
exposure by inhalation and dermal absorption (see below).
Animal studies do indicate that animals exposed to PCB aerosols
show rapid increases in liver PCB levels. Exposure to Pydranl
A 200 for 15 minutes resulted in the accumulation in the
liver of 50 percent of the PCBs accumulated after two hours
(Benthe, et al. 1972). The lung appears to be a good site
of absorption and certain occupational environments contain
significant levels of airborne PCBs. The National Institute
of Occupational Safety and Health has recently proposed
an occupational exposure limit of 1.0 jug/m on a time weighted
average 10-hour day, 40-hour week basis (Natl. Inst. Occup.
Safety Health, 1977). Assuming a tidal air volume of 10
m in an eight-hour day and 100 percent absorption, the
resulting dose at this exposure level would be 10 ^ug/day.
Dermal
Dermal exposure, like inhalation exposure, is a partic-
ularly significant route in the occupational setting. With
C-22
-------
the restriction of PCS uses to sealed systems, the use of
PCBs in products to which the public might be exposed has
declined markedly, reducing opportunities for general exposure.
Past uses of PCBs in carbonless copy paper, printers inks,
and other products probably contributed to general PCS ex-
posures. Documented exposures are largely occupational
as exemplified by the results of Ouw, et al. (1976). The
authors noted that one group of employees were largely ex-
posed through skin contact and had significantly elevated
blood PCB levels.
In a variety of animal studies dermal application of
several PCB containing materials has produced both local
and systemic effects including liver degeneration and death
(Miller, 1944; Paribok, 1954; Vos and Beems, 1971). In
neonatal rats treated by skin application with PCBs, a five-
to ten-fold increase in aryl hydrocarbonhydroxylase activity
occurred in liver, skin, lung, and kidney, indicating signi-
ficant distribution to these tissues after exposure by this
route (Bickers/ 1976; Bickers, et al. 1975).
The relative contributions of various routes of exposure
can be expected to vary widely. Occupational exposures
are by far the most severe with inhalation and skin contact
being the major routes of absorption. A noteworthy by-product
of occupational PCB exposure is the elevated risk of exposure
among other members of workers' families. An epidemiological
study in Bloomington, Indiana revealed significantly elevated
serum PCB levels among a group of 18 occupationally exposed
workers (mean 71.7 ppb) and a slight elevation among 19
C-23
-------
members of their families (near 33.6 ppb) as compared to
background levels (5 to 20 ppb) (McCloskey, et al. 1978).
The general public is widely exposed to PCBs but at much
lower levels and primarily through the diet. Fish living
in contaminated waters are by far the largest contributors
to dietary PCBs (Jelinek and Corneliussen, 1976).
PHARMACOKINETICS
Absorption
The efficiency of PCB absorption in the gut of rats
was shown to be between 92 to 98.9 percent, (Albro and Fishbein,
1972). Neither degree of chlorination (mono-hexachloro-
biphenyl) nor the dose ingested (5 to 100 mg/kg) markedly
affected the efficiency of uptake. Matthews and Anderson
(1975b) observed a reduced accumulation of PCBs in adipose
tissues of rats exposed orally as compared to i.v. injection.
The differences were more pronounced with biphenyls of low
chlorine content and were thought to be related to route
\
of absorption and metabolic rates, rather than to the overall
efficiency of transport across the gut. Absorption via
the gut was also very efficient in adult Rhesus monkeys,
90 percent of a single dose of 1.5 or 3.0 g/kg Aroclor 1248
being absorbed from the gastrointestinal tract (Allen, et
L
al. 1974) .
Efficient absorption via inhalation has been demons-
i.-
trated in rats by Benthe, et al. (1972).
In humans, absorption via the intestine has been best
illustrated by the "Yusho" incident in Japan in 1968. Among
individuals ingesting less than 720 ml of contaminated rfce
C-24
-------
bran oil (equivalent to 1.5 to 2.2 g Kanechlor 400) 39 percent
developed severe symptoms and an additional 49 percent devel-
oped moderate symptoms of PCB intoxication. The lowest
level of PCB ingestion in an affected individual was estimated
to be 0.5 g (Kuratsune, et al. 1972). Absorption via the
respiratory tract and skin is also efficient as indicated
by occupational exposures where effects of PCB exposure
can be detected even at doses too low to produce pathology
(Alvares, et al. 1977).
Distribution
PCBs given to rats by i.v. injection are removed from
the blood rapidly and stored initially in the liver and
muscle. With time they are redistributed primarily to skin
and adipose tissue (Matthews and Anderson, 1975b). The
degree to which PCB's are stored or excreted depends on
their susceptibility to metabolism and, therefore, on the
degree of chlorination and availability of adjacent unsub-
stituted carbons. Tissue levels of mono-, di-, penta- and
hexachloro-biphenyls in rats given a single injected dose
at 0.6 mg/kg were determined by Matthews and Anderson (1975b).
The maximum doses accumulated in each tissue increased with
degree of chlorination as did the half life in each tissue.
The proportion of total PCBs present in tissues as metabolites
was greatest for the mono-and di-chlorobiphenyls. Hexachloro-
biphenyls in tissues were largely unmetabolized. The distri-
bution of PCBs in adipose tissue provides a useful example
of the relative accumulation of different isomers. Tissues
were examined for up to 42 days and a summary of .the results
is presented in Table 6.
C-25
-------
TABLE 6
Storage of PCBs in Adipose Tissue in Rats
(Values are Percent of Total Dose 0.6 mg/kg)
Degree of
Chlorination
mono-
di-
penta-
hexa-
Maximum
11.
52.
23.
85.
63
75
54
18
± 5.
± 14-
± 3-
+ 21.
64
99
0
6
Time of
Maximum
Stored
1
2
1
42
hr
hr
day
days
Amo u n t at
7 Days
0.
1.
13.
56.
234
837
04
08
± °-
± °-
± 2-
+ 15.
055
213
1
72
Adapted from Matthews and Anderson, 1975b.
C-26
-------
A similar pattern was observed in skin with up to 22
percent of the hexachlorobiphenyl dose being accumulated
there at 1 day and residual levels around 15 percent remaining
at 42 days.
Single intravenous doses of 0.6 or 6.0 mg/kg 2, 4,
5, 2',5' pentachlorobiphenyl were cleared from the blood
in ten minutes and initially deposited in liver and muscle.
They were subsequently translocated to adipose tissue and
skin as depositories (Matthews and Anderson, 1975a).
A single administration of approximately 500 mg/kg
2, 5, 2', 51 tetrachlorobiphenyl to rats resulted in a similar
distribution with adipose, skin, and blood being the signifi-
cant storage depots after 24 hours (Van Miller, et al.
1975) .
The significance of chlorine position as well as number
was addressed in a study of the pharmacokinetics of 3, 5,
3',5' tetrachlorobiphenyl (TCB) by Tuey and Matthews (1977).
The arrangement of chlorines on this molecule results in
the absence of adjacent unsubstituted sites. The pattern
of distribution of the compound following a single i.v.
injection of 0.6 mg/kg was similar to that observed in earlier
studies (Matthews and Anderson, 1975a,b) with adipose tissue
and skin becoming the major long term storage sites. However,
loss of 3, 5, 3',5' TCB was slower than earlier observed
for 2, 4, 5, 2', 5' pentachlorobiphenyl (see penta CB Table
6) with the maximum adipose tissue load reaching 52.9 percent
of total dose on day four and the residual on day seven
>
remaining at 45.4 percent. The distribution of several
C-27
-------
tetrachlorobiphenyl isomers in mice was analyzed by Mizutani,
et al. (1977). In all cases the accumulation of the compound
was greater in the carcass than in the liver. A tendency
for those isomers with adjacent unsubstituted carbons to
be rapidly cleared was observed. 2, 6, 2', 6' TCB was very
rapidly cleared from carcass and liver, and 2, 3, 2',3'
TCB was cleared fairly rapidly. However, 2, 4, 2', 4' TCB
was more resistant to removal than 3, 5, 3', 5' TCB which
might not be anticipated on structural grounds. The half
life in the carcass of the former was 9.2 days but only
2.1 days fbr the latter. The degree of accumulation of
the isomers was assessed by the introduction of an index
referred to as a storage ratio (the daily amount entering
storage/daily oral ingestion). By this measure 3, 5, 31,
51 TCB and 2, 4, 2',4' TCB were similar with indices of
0.7 and 0.6, respectively, while the more readily metabolized
2, 3, 2', 3' TCB had an index of 0.06.
The distribution of 2, 5, 21, 5' TCB in infant Rhesus
monkeys was determined after a single dose of tritiated
TCB (500 mg/kg). At 72 hours the distribution differed
from that in rats in that the label was more widely dispersed
in the monkeys. Blood levels were lower than observed in
rats and the major storage depots were bone marrow, adrenal
and skin. Most of the labelled material was associated
with macromolecules although it was largely extractable
and not covalently bound (Hsu, et al. 1975a).
Distribution of PCBs in the human body has not been
the subject of systematic experimentation. Data available
from general population surveys indicate that general patterns
C-28
-------
of distribution are consistent with those found in other
animals. When detected in the adipose tissue of the general
populace, PCB levels ^are around 1 mg/kg (Yobs, 1972;, Kutz
and Strassman, 1976; and Grant, et al. 1976). Plasma levels
detected in the general populace are two to three orders
of magnitude lower than adipose levels (Finklea, et al.
1972) Similarly, Yusho patients exhibited a 100 to 1000
fold greater concentration in the fat of skin, liver and
in adipose tissue than in plasma. Over several years both
the fat and plasma levels were observed to decline to near
normal levels (Kuratsune, et al. 1976). The PCBs found
in humaniadipose tissues in the U.S. chromatographically
resemble Aroclor 1254 and 1260, suggesting that less chlorinated
isomers found in Aroclor 1248 are preferentially excreted
(Kutz and Strassman, 1976).
Metabolism
The metabolism of PCBs has been studied extensively
in several organisms. A detailed review of PCB metabolism
was written by Sundstrom, et al. (1976a). Rather than
attempt to treat the subject exhaustively this section will
summarize the major characteristics of PCB metabolism which
relate to their distribution, accumulation, toxicity, and
possible mechanisms of carcinogenicity.
the metabolism of PCBs depends on their chlorine content,
and the sites of chlorination on the biphenyl (Sundstrom,
et al. 1976a; Lutz, et al. 1977). While the overall mecha-
nisms of metabolism appear to be similar in most vertebrates
examined, the capacity to metabolize PCBs declines from
mammals to birds to fish (Hutzinger, et al. 1972). Elucida-
C-29
-------
tion of PCB metabolism has been made possible by the use
of individual purified isomers. Predominantly, the products
of PCB metabolism at all levels of chlorination are biphenylols,
biphenyldiols, and dihydrodihydroxybiphenyls, although the
types and proportions of specific metabolites vary in different
species. A few biphenyltriols and methoxy derivatives have
also been observed (Sundstrom, et al. 1976a).
The structures of several PCB metabolites support the
formation of arene oxides as intermediates. The first evidence
for the formation of arene oxide intermediates was obtained
by Gardener, et al. (1973). They isolated trans 3, 4 dihy-
droxy-3, 4-dihydro-2, 2', 5, 5' tetrachlorobiphenyl as a
metabolite of 2, 21, 5, 5' tetrachlorobiphenyl in rabbits.
More direct evidence for the formation of arene oxides was
obtained by Safe, et al. (1975, 1976). In rabbits and frogs
the biohydroxylation of 4-chlorobiphenyl was investigated
2
using 4'- H-4-chlorobiphenyl. The major metabolite 4' chloro-
4-biphenylol retained 79 percent of the label which is con-
sistent with arene oxide formation (Daly, et al. 1972).
The subsequent isomerization of the arene oxide results
in the migration of the deuterium atom from the ultimate
site of hydroxylation to the adjacent carbon, an NIH shift.
Daly, et al. (1972) consider the NIH shift of labeled hydro-
gens, halogens or alkyl substituents to be indicative of
enzymatic arene oxide formation. A subsequent hydroxylation
to 4'chloro-3, 4-biphenyldiol resulted in the loss of half
the remaining deuterium suggesting a direct hydroxylation
rather than a second arene oxide formation (Safe, et al.
1975). 4, 4'-dichlorobiphenyl produced three metabolites
C-30
-------
in the rabbit: 4, 4'-dichloro- 3-biphenylol, 3, 4'-dichloro
- 4-biphenylol and 4'-chloro-4-biphenylol. These products
\
are consistent with a mechanism involving 3, 4-arene oxide
formation followed by epoxide ring opening. Either a 1.2
halogen shift, with or without halogen elimination upon
tautomerization, or 3-ol formation after arene ring cleavage
would produce the ultimate products (Safe, et al. 1976;
Safe, et al. unpublished, quoted in Sundstrom, et al. 1976a).
The reactions are diagrammed in Figure 2. Other examples
of PCBs for which metabolic pathways are consistent with
are;ne oxide formation include 2, 2', 4, 4', 5, 5' -hexachloro-
biphenyl in rabbits (Sundstrom, et al. 1976b) and 4 chlorobi-
phenyl and 4, 4'-dichlorobiphenyl in rats (Hass, et al.
1977). Infant Rhesus monkeys fed 2, 5, 2', 5 tetrachloro-
biphenyl excreted dihydroxy, dihydrodihydroxy and dihydrotri-
hydroxy derivatives in urine (Hsu, et al. 1975b).
The K region epoxides of polyaromatic hydrocarbons
are known to bind to nucleic acids in vitro (Grover and
Sims, 1970) and in cultured mammalian cells (Grover, et
al. 1975). Furthermore, they are capable of transforming
cells in culture (Huberman, et al. 1972) although their
significance in tumor induction in animals is in doubt (Grover,
et al. 1975). It has been suggested that arene oxide metabo-
r\
lites of PCBs may react with nucleophilic sites in DNA and
other macromolecules and that alkylation of critical sites
may be involved in the induction of tumors (Allen and Norback
1977).
C-31
-------
urine 2.0
0.2
2.2
2.0
7.5
Figure 2: Metabolic pathways for 4, 4'-dichlorobiphenyl
in the rabbit.
From: Sundstrom, et al. (1976a)
-------
Excretion
The primary routes of PCB excretion are bile (observed
in feces) and urine. Excretion is closely coupled to meta-
bolism. In rats less than ten percent of excreted PCBs
*
were unmetabolized (Matthews and Anderson, 1975b). The
rate and efficiency of excretion were highly dependent upon
the degree of chlorination and structure. Urinary excretion
of PCBs accounted for the removal of 59.8, 33.9, 7.6, and
0.7 percent of total dose of mono, di, penta, and hexachloro-
biphenyl respectively. Over 60 percent of urinary excretion
occurred within the first 24 hours and all urinary excretion
ceased by the ninth and fourth days, respectively, for penta-
and hexachlorobiphenyl (Matthews and Anderson, 1975b).
All the 2, 4, 5, 2',5' pentachlorobiphenyl excreted in urine
by rats was in the form of a glucuronide conjugate of a
metabolite (Chen and Matthews, 1974). While urinary excretion
usually ceases within a few days, biliary excretion continues
for an extended period. The relative contribution of biliary
excretion to the elimination of PCBs increases with chlori-
nation. The kinetics of excretion of mono- and dichlorobi-
phenyl are monophasic while the elimination of penta- and
hexabiphenyl is biphasic. While 90 percent of PCBs up to
pentachlorobiphenyl were excreted in 42 days or less, hexa-
chlorobiphenyl was largely retained in the tissues of the
animal. Extrapolation of the excretion data indicated that
only 20 percent of 2, 4, 5, 21, 41, 5' hexachlorobiphenyl
would even be excreted (Matthews and Anderson, 1975b).
C-33
-------
The- absence of adjacent unsubstituted carbons greatly decreased
excretion as would be expected from the effects of structure
on storage arid metabolism. 3, 5, 31, 51 TCB is excreted
at about the same rate as 2, 4, 5, 2', 5' pentabiphenyl
(Tuey and Matthews, 1977; Matthews and Anderson, 1975a).
While the half-life in fat for 2, 5, 2', 5' TCB was about
33 hours at 500 mg/kg dose in rats (Van Miller, et al.
1975) the half-life for 3, 5, 3', 5' TCB was 12 to 15 days
at dose levels of .6 mg/kg in rats (Tuey and Matthews, 1977).
The half-lives of the individual PCB isomers in the
rat may be approximated by the fecal half-lives which are
15.7 and 22.2 hours for mono and dichlorobiphenyl respectively.
Penta and hexabiphenyls elimination is biphasic with first
and second component half-lives of 39.2 and 211 hours for
penta-CB and 49 and 642 hours for hexa-CB (Anderson, et
al. 1977). Because only 20 percent of the hexa-CB is ulti-
mately excreted its half-life is indefinite.
Rates of elimination of a series of tetrachlorobiphenyls
in mice were determined by Mizutani, et al. (1977). Half-
lives for TCB isomers in liver and the carcass ranged from
0.9 days for 2, 3, 2', 3' TCB to 9.2 and 7.8 days for the
loss of 2, 4, 2', 41 from carcass and liver, respectively.
Structure did not influence elimination as markedly as in
the rat. 3, 5, 3', 5' TCB had half-lives of 2.1 and 2.2
days in carcass and liver. However, stimulation of meta-
bolism by the addition of phenobarbitol did increase the
rate of elimination of 2, 4, 2', 4' TCB more than 3, 5,
3', 5' TCB. The authors concluded that the rate-limiting
C-34
-------
step in the elimination of the isomers was release from
storage in the tissues of the mouse rather than metabolism.
Two differences between the elimination of 2, 5, 2',
5' TCB in infant Rhesus monkeys and rats may be of interest
in evaluating human metabolism. Single doses of 500 mg/kg
to rats resulted in total elimination of about 76 percent
(66 percent feces, 10 percent urine) in 72 hours (Van Miller,
et al. 1975). In primates only one percent of the same
dose was eliminated in feces and two percent in urine after
72 hours (Hsu, et al. 1975a). In addition, the major excreted
metabolite in rats appeared to be 3-hydroxy TCB while a
dihydrodiol TCB predominated in monkeys (Van Miller, et
al. 1975; Hsu, et al. 1975b).
A final comment on the pharmacokinetics of PCBs must
be addressed to transplacental and transmammary movement.
Transplacental uptake of PCBs by a fetus has been documented
in mice (Masuda, et al. 1978), rats (Curley, et al. 1973),
Rhesus monkeys (Allen and Barsotti, 1976), and humans (Yosh-
imura, 1974) . In mice, transplacental and transmammary
uptake of PCBs were approximately 0.1 to 0.2 and 20 to 35
percent of total dose respectively (Masuda, et al. 1978).
Similar values were observed in rats (Mizunoya, et al. 1974).
Female monkeys consuming 2.5 ppm Aroclor 1254 transferred
enough via breast milk to produce severe hyperplastic gas-
tritis in nursing infants (Allen and Barsotti, 1976). Recently,
a preliminary mathematical model of PCB distribution in
rats has been proposed (Lutz, et al. 1977; Anderson, et
al. 1977).
C-35
-------
It should be noted that most of the laboratory studies
discussed above have been performed with pure isomers, while
toxicity studies and environmental exposures involve commercial
mixtures with possible dibenzofuran contamination. In addition,
commercial mixtures tend to contain asymmetrical polychlorinated
biphenyls (Natl. Inst. Occup. Safety Health, 1977).
The pharmacokinetics of PCBs can be summarized with
the following points:
1. They are readily absorbed through the gut, respiratory
system, and skin.
2. They may initially concentrate in the liver, blood,
and muscle mass; but long-term storage in mammals is
primarily in adipose tissue and skin.
3. The major metabolic products of PCBs are phenolic deriv-
atives or dihydrodiols which may be formed through
pathways with arene oxide intermediates or by direct
hydroxylation. The susceptibility of individual PCB
isomers to metabolism is a function of the number of
chlorines present on the biphenyl and their arrangement.
Biphenyls which have one or more pairs of adjacent
unsubstituted carbons are more rapidly metabolized
than those which do not.
4. PCBs which are readily metabolized are also rapidly
excreted in the urine and bile. Excretion in urine
is most prominent for the least chlorinated, while
bile becomes the more significant route of excretion
for more highly chlorinated isomers.
C-36
-------
5. "Those isomers which are most refractory to metabolism
accumulate for increasing periods of time in fatty
tissues. Highly chlorinated isomers are accumulated
almost indefinitely.
6. PCBs can be transferred either transplacentally or
in breast milk.
7. Non-human primates may retain PCBs more efficiently
than rodents.
EFFECTS
Acute,, Sub-acute, and Chronic Toxicity
Several reviews of the toxic effects of PCBs in animals
and man have appeared in recent years (Kimbrough, 1974;
Fishbein, 1974; Peakall, 1975; Kimbrough, et al. 1978; Cordle,
et al. 1978; Natl. Inst. Occup. Safety Health, 1977 (which
is particularly recommended for human effects)). This section
will attempt to highlight the most significant toxic effects
observed in animals and man, but will not seek to be compre-
hensive.
The acute oral and dermal LD^Q for PCBs in rats, mice,
and rabbits are given in Tables 7, 8, and 9. In the classi-
fication by the American Hygiene Association the PCBs, are
slightly toxic or almost nontoxic (Hodge and Sterner, 1949).
In rats, Bruckner, et al. (1973) observed a 14-day LD5Q
of 4.25 g/kg. Toxic effects of high doses of Aroclor 1242
included diarrhea, chromoacryorrhea, loss of body weight,
unusual stance and gait, lack of response to pain stimuli,
and terminal ataxia. CNS deterioration and dehydrydration
were thought to be contributing factors. Histopathologic
changes were observed only in liver and kidney. Miller
C-37
-------
TABLE 7
Acute Toxicity of PCD's in Several Strains of Rats and Mice.
Compound tested
Species and sex
Route
LD
g/kg body weight
Reference
o
i
U>
00
Aroclor 1254
Aroclor 1260
Aroclor 1254
Aroclor 1260
Aroclor 1254
Aroclor 1221
•Aroclor 1262
Aroclor 1240
Aroclor 1254
Aroclor 1254
Aroclor 1254
Aroclor 1254
Kaneclor-400
Kaneclor-400
Kaneclor-400
Kaneclor-400
Kaneclor-300
Kaneclor-300
nP-200 biphenyls of
dichloride and below
2,4' -Dichlorobiphenyl
Trlchlorobiphenyl
niphenyl of 'trichloride
and below
2,4,3',4'-Tatrachlorobiphenyl
5-OH derivative of 2,4,3',4'-
tetrachlorob-phenyl .
2,3,4,3' ,4' -P-.-atachlorobtphenyl
Rat (adult, Sherman strain)
Rat (adult, Sherman strain)
Rat (weanling, Sherman strain)
Rat (weanling, Sherman strain)
Rat (female, Sherman strain)
Rat (female, Sherman strain)
Rat (female, Sherman strain)
Rat
Rat (Wistar, 30-day-old, M-F)
Rat (Wistar, 60-day-old, M-F)
Rat (Wistar, 120-day-old, M-F)
Rat (Wistar, 120-day-old, F)
Rat (Wistar, M)
Rat (Wistar strain, F)
Mice (CPI strain, M)
Mice (CFI strain, F)
Rat (Wistar strain, M)
Rat (Wistar strain, F)
Mice (dd strain, F)
Mice (dd strain, F)
Mice (dd strain, F)
Mice (dd strain, F)
Mice (DVI strain)
Mice (CFI strain)
Mice (CFI strain)
"•"^afcrence nutnt_ fre**. source.'
•;:.: K'.mbrough, *- «?.. 1578.
Oral
Oral
Oral
Oral
Intravenous
Oral
Oral
Oral
Oral
Oral
Oral
Oral
Oral
Oral
Oral
Oral
Oral
Oral
/ Oral
Oral
Oral
Oral
Intraper itoheal
.Intraperttoneal
Intrapcrttoneal
4-10
4-10
1.295
1.315
0.358
4,00
11.3
4.25
1.3
1.4
2.0
2.5
(ml kg)
(ml kg)
30
14
1.075 (ml kg)
1.57
(ml kg)
,15
.05
6.36
7.86
3.0fi - 4.25
9.27
2.15
0.43
0.65
(5)
(5)
(5)
(5)
(5)
(6)
(6)
(7)
(8)
(8)
(8)
(8)
(9)'
(9)
(9)
(9)
(9)
(9)
(10)
(10)
(10)
(10)
(11)
(11)
(11)
-------
TABLE 8
a
Oral LD5Q (rat)
LD
Compound tested g/kg body weight
Aroclor
Aroclor
Aroclor
Aroelor
Aroclor
Aroclor
Aroclor
1221
1232
1242
1248
1260
1262
1268
(Und
(Und
iluted)
iluted)
(Undiluted)
(Und
(50%
(50%
(33.
iluted)
soln in
soln in
3% soln
corn
corn
in
oil)
oil)
corn
oil)
2.
1.
0.
0.
1.
1.
2.
000
26
794
794
26
26
5
- 3.
- 2.
- 1.
- 1.
- 2.
- 3.
169
0
269
269
0
16
aData of Panel on Hazardous Substances (6)
From: Kimbrough, et al. (1978)
TABLE 9
a
Skin LD5Q (rabbits)
LD
Compound tested g/kg body weight
Aroclor
Aroclor
Aroclor
Aroclor
Aroclor
Aroclor
Aroclor
1221
1232
1242
1248
1260
1262
1268
(Und
(Und
(Und
(Und
(50%
(50%
(50%
iluted
iluted
iluted
iluted
soln
soln
soln
,
)
)
)
in
in
in
corn oi
corn oi
corn oi
1)
1)
1)
3.
4.
8.
11.
10.
11.
10.
98
47
65
0
0
3
9
aData of Panel on Hazardous Substances (6)
From: Kimbrough, et al. (1978)
C-39
-------
(1944) found the guinea pig most sensitive to Aroclor 1242
followed by the rabbit and rat. In the rat, toxicity de-
creased with increasing degree of chlorination; however,
the effect was not observed with rabbits (Fishbein, 1972).
The more significant toxic effects of PCBs are
observed on repeated exposure over a period of time. Aroclor
1254 at 1000 ppm in the diet was fatal to 75 percent of
male rats in 43 days with total intakes of 500 to 2000 mg/kg
being lethal (Tucker and Crabtree, 1970). Phenoclor DP6 "
fed at 2000 ppm to rats resulted in marked weight loss and
death between 12 and 56 days after the initiation of treatment
(Vos and Koeman, 1970). Guinea pigs treated dermally for
11 days with a total of 379.5 mg of a PCB with 42 percent
average chlorine content died at intervals up to 21 days
following the first application (Miller, 1944). Aroclor
1254 at 1000 ppm in the diet killed 5/10 male rats and 8/10
female rats. At 500 ppm over eight months two males and
one female died while no lethality was observed at 100 or
20 ppm. Aroclor 1260 was less toxic, with 8/10 females but
no males dying at 1000 ppm. No males died at lower doses
and 1/10 and 2/10 females died at 100 and 500 ppm respec-
tively. Substantial weight losses were observed at 100
and 500 ppm in both males and females (Kimbrough, et al.
1972). Mink have been shown to be unusually sensitive to
PCBs. A mixture of Aroclors 1242, 1248 and 1254 at 30 ppm
in the diet for 6 months was 100 percent lethal (Aulerich,
et al. 1973) as was 3.6 ppm Aroclor 1254 over 105 days
in another study (Plantonow and Karstad, 1973). Adult Rhesus
monkeys (Macaca mulatta) were particularly sensitive to
C-40
-------
PCBs. Aroclor 1248 at 100 or 300 ppm in the diet for two
to three months resulted in extreme morbidity within one
month and almost 100 percent mortality within three months.
Total intakes for the two groups were 0.8 to 1.0 g for 100
ppm and 3.6 to 5.4 g for 300 ppm (Allen, 1975).
The roost consistent pathological changes occurring
in mammals after PCB exposure are in the liver. In rats/
i
rabbits, and guinea pigs, Miller (1944) observed fatty depo-
sits after acute injections and similar changes in rabbits
and guinea pigs after dermal application. In feeding experi-
ments, marked fatty metamorphosis was noted in guinea pig
liver with intracellular hyaline bodies being observed in
rats. Less striking changes were noted in the kidneys,
lungs, adrenals, and heart of guinea pigs. Rats exposed
repeatedly ;to dietary PCBs show increased liver weights
(Kimbrough, et al. 1972; Bruckner, et al. 1973). Kimbrough,
et al. (1972) fed rats Aroclor 1254 or 1260 at levels between
20 and 1000 ppm for eight months. Light microscopic changes
observed included hypertrophy of liver cells, cytoplasmic
inclusions, brown pigment in Kupffer cells, lipid accumula-
tion and, at higher doses, adenofibrosis. Ultrastructural
examination revealed an increase in smooth endoplasmic retic-
ulum. The effect of Aroclor 1254 was more pronounced than
that of 1260. Porphyria was observed in the livers and,
occasionally, other tissues of animals exposed to either
mixture.
C-41
-------
Rats fed 2000 ppm Phenoclor DP6 also had enlarged livers
with vacuolated foamy cells containing pycnotic nuclei (Vos
and Koeman, 1970). Vacuolization of liver cells was also
noted by Bruckner, et al. (1973) after dosing rats with
100 rag/kg Aroclor for three weeks although no overt toxicity
was manifest.
Rats fed 100 ppm Aroclor 1242 (6.6 to 3.89 mg/kg/day)
or Aroclor 1016 (6.9 to 3.5 mg/kg/day) for periods of up
to ten months showed no signs of overt intoxication or gross
liver changes. Enlarged hepatocytes with vacuolated cyto-
plasms and inclusions were noted. Aroclor 1242 seemed to
produce more pronounced changes than 1016. Four and six
months after the discontinuation of exposure hepatocytes
were still enlarged but cytoplasmic vacuoles and inclusions
had diminished, suggesting a degree of reversibility of
effect. Significant residual levels of PCBs remained in
adipose tissue. Using electron microscopy, increased smooth
endoplasmic reticulum and lipid vacuoles as well as atypical
mitochondria were observed. No significant gross changes
in other organs were noted (Burse, et al. 1974).
Allen and Abrahamson (1973) fed rats 1000 ppm of either
Aroclor 1248, 1254, or 1262 for 1, 3, 7, 14, 21, or 28 days
or 6 weeks. No overt toxicity was observed although weight
gain was retarded in all treated groups. The effect was
inversely proportional to percent chlorination. Increased
liver size, protein, and RNA content were observed. The
magnitude of changes increased with the percent chlorination.
Hypertrophy was associated with proliferation of the smooth
endoplasmic reticulum, formation of membranous arrays, and
increased lipid droplets.
C-42
-------
The effect of metabolism on toxicity was explored •
by giving rats large (1.5 g/kg) single doses of 2, 5, 2',
S'-tetrachlorobiphenyls which produced high mortality within
two to three days (Allen, et al. 1975). Pretreatment with
phenobarbitol to induce metabolic enzymes allowed survival
without obvious ill effects following a 1.25 g/kg dose,
while treatment with the microsomal enzyme inhibitor SKF
525A lead to 100 percent mortality in four days. The ability
to metabolize and eliminate TCB appears to protect the animal.
Dietary administration of 100 ppm TCB for three weeks produced
less liver hypertrophy than Aroclor 1248.
Liver pathology in mice exposed to 1.5 mg PCB/day was
essentailly the same as seen in rats/ including increased
smooth endoplasmic reticulum and increased lipid droplets
(Nishizumi, 1970).
Rabbits receiving 300 mg orally of Aroclor 1221, 1242,
or 1254 for 14 weeks were examined (Roller and Zinkl, 1973).
Aroclor 1221 and 1242 treated rabbits gained weight at control
rates while 1254 treated rabbits did not gain as much.
Livers of 1254 and 1242 treated animals were enlarged while
livers of 1221 treated animals were smaller than controls.
Gross liver lesions and small uteri were apparent in the
1254 treated animals but not the others. Liver pathology
in 1254 treated animals included enlarged hepatocytes with
foamy to granular cytoplasms, and subcapsular midzonal necrosis
Aroclor 1242 produced a liver pathology similar to 1254.
Aroclor 1221 treated animals were free of histologic changes.
C-43
-------
Dermal studies with rabbits using Clophen A60, Phenoclor
DP6 and Aroclor 1260 indicated that the latter was the least
toxic (Vos and Beems, 1971). The former two mixtures had
been shown to be contaminated with tetra- and penta-chlorodi-
benzofuran (Vos, et al. 1970). Skin lesions produced included
hyperplasia and hyperkeratosis of the epidermal and follicular
epithelium, and were accompanied by pathological changes
in the liver and kidney. The chlorodibehzofuran impurities
in the PCBs were thought to be responsible for the skin
lesions. A comparison of the toxic effects of dermally
applied 2, 4, 5, 2'f 5' hexachlorobiphenyl and Aroclor 1260
demonstrated that the skin lesions appeared sooner and were
more severe after treatment with the commercial mixture.
Liver changes were found in both treatment groups with the
pure isomer inducing the more severe effects. From this
study it was concluded that the chlorodibenzofuran contaminants
in commercial mixtures probably contribute to the skin lesions
(chloracne), edema formation, and liver damage while PCBs
contribute in lesser degrees to chloracne and liver damage
but are primarily responsible for the hepatic porphyria
observed in PCB intoxication (Vos and Notenboom-Ram, 1972).
Non-human primates are rather sensitive to PCBs. Male
Rhesus monkeys were fed 300 ppm Aroclor 1248 for three months.
Effects which began to appear within a month included hair
loss, subcutaneous edema, purulent discharge from the eyes,
acneform eruptions, and liver hypertrophy caused by smooth
endoplasmic reticulum proliferation. Marked hypertrophy
of the gastric mucosa was a significant finding not usually
C-44
-------
seen in rodents. Invasion of the submucosa by the mucosal
epithelium with increased cellularity and dysplasia occurred
in the stomach. The dietary levels used were about tenfold
greater than the contamination levels in foods during the
early 1970's and the gastric changes observed were considered
to be of particular significance to human risk (Allen and
Norback, 1973). When fed low levels (2.5 and 5 ppm) of
Aroclor 1248 for 52 weeks female monkeys developed periorbital
edema, alopecia, erythema and acneform lesions. Effects
in males were less pronounced (Barsotti and Allen, 1975) .
The high sensitivity of monkeys to PCBs has been confirmed
and the evaluation of the toxic effects, particularly in
the gastric mucosa, has been extended (McNulty, 1976; Bell,
1976) . The pathologic effects of PCBs in nonhuman primates
have been reviewed by Allen and Norback (1976) and Allen
(1975).
The ability of PCBs to induce liver microsomal enzymes
was demonstrated by Street, et al. (1969). Enzyme induction
by commercial PCBs has been shown in rabbits (Villeneuve,
et al. 1971a), rats (Litterst and VanLoon, 1972), and primates
(Allen, et al. 1974). In rats induction is observed following
intraperitoneal injection (Bickers, et al. 1972) or skin
application (Bickers, et al. 1975). Threshold values for
enzyme induction vary between 0.5 and 25 ppm (Villenueve,
et al. 1971a; Litterst, et al. 1972; Turner and Green, 1974) .
The induction of demethylating activity in rats by Aroclor
1254 was maximum in seven days while cytochrome P450 and
nitroreductase activities continued to rise over four weeks
C-45
-------
of treatment. Activities declined slowly after discontinua-
tion of treatment reaching control levels in about ten days
(Litterst and VanLoon, 1974). Cutaneous exposure to PCBs
resulted in a maximum induction within two to six days.
(Bickers, et al. 1972, 1975). Degree of induction of enzyme
activities was found to correspond to increasing chlorine
content of Aroclors (Litterst, et al. 1972) and of di, tetra,
and hexachlorobiphenyl mixtures (Schmoldt, et al. 1974).
The effects of chlorine content and position of pure isomers
were examined by Johnstone, et al. (1974), Ecobichon (1975),
and Ecobichon and Comeau (1975). More highly chlorinated
isomers and those substituted at the 4 and 4' positions
were most active in inducing enzymes associated with the
endoplasmic reticulum. For less localized enzymes, position
was less critical, although chlorinated compounds were more
effective than biphenyl.
The effects of dietary exposure to Aroclor 1254 on
enzyme induction were investigated in rats by Bruckner,
et al. (1977). Aroclor 1254 at 5 or 25 ppm induced dose
dependent increases in the metabolism of pentobarbitol,
aminopyrine, and acetanilide after 35, 70, and 140 days
of exposure. Exposure to 1 ppm had little effect on metabolism.
Liver weight and serum triglyceride levels were elevated
only in animals exposed to 25 ppm. In 15-day experiments
induction of aminopyrine N-demethylation was observed after
the first day of exposure at 5 and 25 ppm, and acetanilide
hydroxylation was induced after two days. Aminopyrine N-
demethylation returned to normal 15 days after the termination
of exposure. Consumption of as little as 1 to 2 mg of PCBs
in 24 hours was sufficient to stimulate acetanilide hydroxylation
C-46
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Commercial PCBs have been shown to induce cytochrome
P450 (phenobarbitol type) and cytochrome P448 (3 methylcho-
lanthrene type) (Alvares, et al. 1973). More recent studies
with purified isomers indicated that ortho-para-substituted
PCBs induce P450 while meta-para-substituted PCBs induce
P448. Substitution in the ortho-position dominates over
meta and no isomers were found to induce both activities
(Goldstein, et al. 1977). The induction of both systems
by commercial preparations and some purified isomers has
recently been shown to result from contamination with dibenzo-
furans. Even "99 percent pure" isomeric PCBs containing
44 ppm tetrachlorodibenzofuran effectively induces P448
while more rigorously purified material does not (Goldstein,
et al. 1978). This observation serves as a reminder that
the effects of trace contaminants must be kept in mind when
evaluating the toxic effects of PCBs.
Enzyme inducing effects of PCBs have also been examined
in vivo by the observation of shortened phenobarbitol sleeping
times in PCB treated animals (Bickers,et al. 1972; Johnstone,
et al. 1974; and Villeneuve, et al. 1972). PCB induction
of enzyme activities in other tissues has included skin
(Bickers, et al. 1975) placenta and fetus (Alvares and Kappas,
1975), neonatal liver during lactation (Alvares and Kappas,
1975), and lung and kidney (Vainio, 1974) .
Other systemic effects of PCBs in mammals include por-
phyria (Bruckner, et al. 1974), increased thyroxin metabolism
(Bastomsky, 1974) and ultrastructural changes in the thyroid
(Collins, et al. 1977), inhibition of ATPases (LaRocca and
c-47
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Carlson, 1975), and interference with oxidative phosphorylation
(Sivalingan, et al. 1973). Alterations in steroid hormone
metabolism are produced by PCBs in rats (Hitman and Cecil,
1970), mice (Orberg and Kihlstrom, 1973), and other animals.
Aroclor 1254 appears to reduce liver vitamin A concentrations
in pregnant rabbits (Villeneuve, et al. 1971b). A more
complete review of these effects can be found in Matthews,
et al. (1978).
PCBs have been shown to have immunosuppressive effects
in rabbits (Vos and Beems, 1971; Street and Sharma, 1975),
guinea pigs (Vos and van Genderen, 1973; Vos and DeRoij,
1972), monkeys, mice (Thomas and Hinsdill, 1978) , and several
birds. Significant effects were observed in Rhesus monkeys
exposed to dietary levels of Aroclor 1248 as low as 5.0
ppm.
Effects of Aroclor 1254 and 1260 on reproduction in
Sherman strain rats were investigated (Linder, et al. 1974).
Dietary levels of 5 ppm Aroclor 1254 had no effect on repro-
duction in rats exposed through two generations. Liver
weights were increased in male and female offspring of the
F, and F2 generations. At 1 ppm, Aroclor 1254 caused increased
liver weights in F-^ male weanlings. At 20 ppm Aroclor 1254
the number of pups in the F-^^ and F2 generations was reduced
while 100 ppm resulted in increased mortality in F-,b offspring
and decreased the mating performance of F-,b adults. Aroclor
1260 produced increased liver weights in Fj_ offspring at
5 ppm but did not affect reproduction at 100 ppm. At 500
ppm litter sizes were reduced and survival was decreased
C-48
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in F, litters. Pregnant rats given 100 mg/kg/day Aroclor
1254 on days 7 to 15 had grossly normal litters but only 30.1
percent survived to weaning. Dosage rates of 50 mg/kg/day
Aroclor 1254 or 100 mg/kg/day Aroclor 1260 did not affect
reproduction or pup survival.
Rabbits fed 0.1 or 1.0 rag/kg body weight Aroclors 1221
or 1254 showed no significant decrease in number of pregnancies
or number of fetuses per litter (Villeneuve, et al. 1971a).
No induction of fetal liver enzymes could be detected.
However, administration during gestation of 600-2, 500 ppm
Aroclor 1254 in the diet resulted in resorptions, abortions,
maternal death and, in two fetuses, asymmetric skulls (Ville-
neuve, et al. 1971b).
Reproductive effects in mice were investigated in animals
treated for ten weeks with 0.025 mg/day Clophen A60 (Orberg
and Kihlstrom, 1973). The length of the estrus cycle was
increased from 6.6 days in controls to 8.7 days in experimental
animals. Also the percentage of implanted ova was reduced
from 87.0 to 79.5. In a second study the reproductive effects
of neonatal exposure to PCBs in milk were examined by injecting
lactating female mice with Clophen A60. On the day of partu-
rition and at weekly intervals for three weeks, the females
were injected with 50 mg of PCB. When treated male and
female offspring were mated with each other, the percent
implantation dropped from a control level of 94 percent
to 75 percent (Kihlstrom, et al. 1975).
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In female Rhesus monkeys exposure to 25 ppm Aroclor
1248 in the diet for two months lead to the typical effects
of PCB intoxication for monkeys including edema, alopecia,
and acne. One animal ingesting a total of 450 mg PCB died
two months after exposure ended and was found to have hyper-
plastic gastritis and bone marrow hypoplasia. The remaining
five animals were bred three months after treatment. Three
were thought to have conceived but resorbed or aborted the
embryos in the first two months of pregnancy. One delivered
a fully developed but small infant (Allen, et al. 1974).
In a more fully developed study both male and female
Rhesus monkeys were fed either 2.5 or 5.0 ppm Aroclor 1248
in the diet (Barsotti and Allen, 1975; Barsotti, et al.
1976) . The total intake in the first 6 months for the females
was 180 and 364 mg for the 2.5 and 5.0 ppm diets, respectively.
Untreated females bred to treated males had normal rates
of conception (Barsotti and Allen, 1975). Treated females
bred to normal males produced the following rates of conception:
control, 12/12; 2.5 ppm, 8/8; 5.0 ppm, 6/8. Live births
resulting from the conceptions were: control, 12/12; 2.5
ppm, 5/8; 5.0 ppm, 1/6. In the 2.5 ppm group, three fetuses
were resorbed shortly after conception. In 5.0 ppm group
three pregnancies aborted at 46, 67, and 107 days of gestation,
one fetus was resorbed, one was stillborn, and one normal
birth occurred. The two females who failed to conceive
were subsequently bred five times without conception. The
live born infants were of low birth weight and showed signs
of PCB intoxication after nursing their mothers for less
C-50
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than two months. Three infants died 44 to 112 days after
birth (Barsotti, et al. 1976). The mothers' breast milk
contained 0.154 to 0.397 ppm PCBs and one contained 16.44
ppm (fat basis) (Allen and Barsotti, 1976). It should be
noted that the dose levels producing these rather striking
effects are within the range of contamination of the human
diet observed until the mid 1970's.
Recently, adipose tissue levels of PCBs in infant Rhesus
monkeys exposed in utero and via breast milk have been corre-
lated with behavioral effects (Bowman, et al. 1978). Three
of five infants born to mothers exposed to 2.5 ppm Aroclor
1248 in the diet during pregnancy and lactation survived
over four months. PCB levels in fat tissue in the infants
declined with a first order rate constant over a period
of 8 to 23 months of age. Extrapolated maximum PCB levels
were 21, 114, and 123 ug/g fat. A battery of eleven behavioral
tests was conducted with the three exposed animals and four
controls over this time period and a positive correlation
between reduced performance and PCB body burden was observed
for seven tests.
Mink have been found to be exceedingly sensitive to
PCB-induced reproductive failure. A marked increase in
kit mortality was observed in commercial mink in the mid-
1960 's after fish meal derived from spawning Great Lakes
Coho salmon was incorporated into the diet. Laboratory
studies confirmed that the reproductive losses were related
to the ingestion of Great Lakes fish (Aulerich, et al. 1971)
and subsequent investigation showed that PCBs contaminating
C-51
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the fish meal were the probable toxic agents (Ringer, et
al. 1972). When fed 10 ppm each of Aroclors 1242, 1248,
and 1254 (30 ppm total) 11/11 adult female mink died prior
to the end of the normal whelping (delivery) period (Ringer,
et al. 1972). Aroclor 1254 fed at 10 ppm resulted in no
offspring among six females. At 5 ppm, Aroclor 1254 fed
for four months prior to whelping depressed reproduction
with only 3 of 12 females whelping and 3 of 9 kits born
alive. At 1 ppm Aroclor 1254, 8 of 10 females whelped and
35 of 43 kits were born alive. Among control animals 11
of 11 whelped and 56 of 66 pups were alive at birth. The
reproductive toxicity of Aroclor 1254 becomes pronounced
between 1 and 5 ppm in the diet (Ringer, et al. 1972).
At 2 ppm in a nine month feeding trial, Aroclor 1254 signifi-
cantly reduced reproduction while Aroclors 1016, 1221, and
1242 did not (Aulerich and Ringer, 1977). Assuming a food
intake of 150 gm/day (Schaible, 1970) the total PCB intake
in the two trials would have been 90 mg at 5 ppm for four
months or 61 mg at 2 ppm for nine months (Aulerich and Ringer,
1977) .
Human exposures to PCBs resulting in toxic effects
have almost all resulted from the ingestion of rice oil
contaminated with Kanechlor 400 in Japan or from industrial
exposure. While absorption through the gut was the route
of exposure in the former case, occupational exposures occur
largely by inhalation or absorption through the skin.
Yusho, the disease resulting from the ingestion of
contaminated rice oil in Japan, has been the subject of
continuing study since the episode of exposure in 1968.
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Periodically, special reports on these continuing studies
have been published in Fukuoka Acta Med (Vol. 60, 1969;
Vol. 62, 1971; Vol. 63, 1972; Vol. 65, 1974; Vol 66, 1975;
Vol. 68, 1977). These results, largely published in Japanese,
have been reviewed in English by the Japanese investigators
both early in the study (Kuratsune, et al. 1972; Kuratsune,
1972) and more recently (Kuratsune, et al. 1976). The cause
l
and scope of the exposure of the Japanese public has been
described above (See Ingestion from Food). The initial
symptoms of Yusho included increased eye discharge and swelling
of upper eyelids, acne-form eruptions and follicular accentua-
tion, and pigmentation of the skj.n. Other symptoms including
dermatologic problems, swelling, jaundice, numbness of limbs,
spasms, hearing and vision problems, and gastrointestinal
disturbances were prominent among the complaints of patients
seen within the first eight months after exposure (Kuratsune,
et al. 1972). The first patients were seen almost immediately
after the release of the contaminated oil in February 1968.
Of a group of patients seen between October 1968 and January
1969, 55 percent became ill between June and August. It
was ultimately determined that as many as 63.9 percent of
those who consumed contaminated oil became ill. Among a
group of 146 known users of the oil, 80 consumed less than
720 ml and 88 percent of these users were affected. Among
those who used more than 720 ml, 100 percent were affected.
The clinical severity of symptoms did not differ by sex
but the age group 13 to 29 was more affected than others
(Kuratsune, et al. 1972).
C-53
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The analysis of the oil indicated that it contained
between 2 and 3 mg/kg of Kanechlor 400 (Kuratsune, et al.
1972). It was later discovered that Kanechlor 400 contained
18 ppm of polychlorinated dibenzofurans (PCDFs) and that
the PCDF concentration in "Yusho Oil" was about 5 ppm .(Nagayama
et al. 1975). The PCDF level in the oil was 250 times greater
than would be expected based on the level in fresh Kanechlor
400, leading Kuratsune, et al. (1976) to suggest that the
concentration increased with PCB use as a heat transfer
med ium.
The amounts of Kanechlor 400 ingested were estimated
for the original 146 person study group. The average amount
ingested was estimated to be 2 g while the minimum amount
ingested by a patient was about 0.5 g (Kuratsune, et al.
1972).
Laboratory evaluations of patients during the early
period were summarized by Kuratsune (1972). Several changes
in blood were noted, including decrease in erythrocyte count,
increase in leukocyte count, and increase in serum lipids,
particularly triglycerides. Blood proteins, electrolytes,
and enzyme activities were normal in most instances. Some
increases in urinary ketosteroid excretion were observed.
The "cheesy" material from Yusho acne contained more steric
and oleic acids than did "normal acne", but less myristic
palmitic and palmitoleic acid. Linoleic acid was present
in Yusho acne but not "normal acne." Liver biopsy indicated
hypertrophy of the smooth endoplasmic reticulum, reduction
of the rough endoplasmic reticulum, filamentous inclusions,
C-54
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and mitochondrial abnormalities. Skin changes included
hyperkeratosis, cystic dilatation of the hair follicles
and marked increase of melanine in basal cells of the epidermis
Decreased sensory nerve conduction velocities were observed
in 9 of 23 patients. Abnormalities of the eyes included
hypersecretion of the meibomian gland, and abnormal pigmenta-
tion of the conjunctiva.
Thirteen women/ 11 with Yusho and 2 without, but married
to men wi'th Yusho, 'delivered ten live and two stillborn
infants between February 15 and December 31, 1968. Nine
of the ten had grayish-dark stained skin, and five had similar
pigmentation of the gingiva and nails. Eye discharge was
common. A stillborn fetus had marked hyperkeratosis, atrophy
of the epidermis, and cystic dilatation of the hair follicle.
Increased melanin pigment in the blood cells and the epidermis
was also noted. Twelve of the 13 fetuses were small for
date of birth. The growth of children affected by Yusho
was significantly lower than Japanese national standards.
A detailed clinical study of four Yusho babies showed that
they were small for their age, had dark pigmentation on
skin and mucous membranes, and gingival hyperplasia. Teeth
were erupted at birth, spotted calcification of the parietooc-
cipital skull, wide fontanels, and saggital suture was present,
along with facial edema and exophthalmic eyes (Yamashita,
1977).
By three years after the episode about half the patients
were improving while 40 percent were essentially unchanged
and 10 percent were becoming more severely affected. Even
C-55
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among those said to be improving, many still complained
of persistant headaches, general fatigue, weakness and numbness
of limbs, weight loss, and other problems (Kuratsune, et
al. 1972).
An evaluation of the longer-term effects of Yusho has
been summarized by Kuratsune, et al. (1976). In 1972 Masuda
noted a peculiar gas chromatographic pattern of PCB fractions
which was common to blood, tissues and breast milk of Yusho
patients (Koda and Masuda, 1975) . A pattern seen in about
60 percent of Yusho patients contained a larger amount of
a late eluting peak than PCB-containing tissues resulting
from other types of exposures. This pattern was referred
to as type A. A similar pattern seen in about 37 percent
of Yusho patients was referred to as type B. These two
patterns (types A and B) have never been observed in individuals
(human or animal) exposed to PCBs in other situations.
These types appear unique to Yusho. Tissue levels of PCBs
in patients undergoing surgery or who died and were autopsied
were followed over several years. Adipose tissue levels
were high (13 to 76 ppm) shortly after the end of exposure
but were substantially lower by the next year. By 1970
and beyond, tissue levels were within the normal range in
the cases studied. Blood levels were not determined until
1972 by which time they were in the normal range. Patients
whose plasma PCB pattern was type A had higher levels than
those with type B.
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The discovery of substantial levels of PCDF in Yusho
oil has been discussed. Levels of PCDFs in control individuals
and Yusho patients were determined. No detectable ( 0.1
ppb) PCDFs were found in controls while tissues of patients
who died in 1969 and 1972 contained .009 and .013 ppm in
adipose and liver respectively. Ratios of PCB/PCDF were
144 and 4 for adipose tissue and liver, respectively. PCDF
levels were higher in liver than adipose on a fat basis.
Although the sample was small, the levels in whole adipose
tissue appeared to have dropped to about 1/3 of the 1969
level by 1972.
By 1972 the dermal and mucosal signs which were most
marked in the initial stages of toxicity were gradually
improving. Symptoms considered to be due to internal distur-
bances, such as fatigue, poor appetite, abdominal pain,
headache, pain and numbness in the limbs, and cough and
expectoration of sputum, have become more prominent. Between
March 1973 and April 1974, 79 patients were examined and
blood PCBs evaluated (Koda and Masuda, 1975). Of patients
with type A or B plasma PCB chromatographic patterns a majority
exhibited some or all of the typical spectrum of dermatological
symptoms with frequencies in type A patients being higher
than in type B patients. Because PCB levels in type A patients
were higher than in type B, the severity of symptoms was
correlated with blood PCB levels.
Serum triglyceride levels in males did not decline
significantly between 1969 and 1974 (Okumura, et al. 1974) .
Levels in female patients declined but were still above
c-57
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normal. The elevation of triglycerides correlated with
increased blood PCB levels and the type A pattern.
Serum bilirubin in patients was lower in 121 patients
than in 257 controls, indicating an accelerated rate of
disposal (Hirayama, et al. 1974) .
Long-term effects continued to be observed in children
born to Yusho mothers,. Nine infants with dark brown skin
pigmentation were born to Yusho mothers between 1969 and
1972, three of them to a patient between 1969 and 1971
(Yoshimura, 1974). The plasma PCB levels of 30 children
born to 18 Yusho mothers were significantly above control
levels but lower than maternal levels (Abe, et al. 1975).
Children nursed by their mothers had higher levels than
children who were not breast fed. One case was reported
by Yoshimura (1974) in which a baby was thought to have
acquired Yusho solely as a result of breast milk intake.
Masuda, et al. (1974) found PCB levels in breast milk
of five Yusho women between 0.03 and 0.06 ppm which was
just within the normal range. A recent study of PCB levels
in the breast milk in 400 Japanese women detected average
levels of 0.033, 0.026, and 0.029 ppm in three measurements
made at two month intervals (Yakushiji, et al. 1977). Based
on these levels, they calculated that daily intake by a
nursing infant would be 24 /jg/day. This can be compared
to an average dietary intake by Japanese adults of 21>ug/day
or 9 jug/day by U.S. adults. By April 30, 1975, 29 of 1,291
Yusho patients had died. Among 22 who died before September,
1973, 9 deaths resulted from malignant neoplasms (Urabe,
1974) .
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The occurrence of Yusho symptoms after modest PCB intake
coupled with the similarity of many of the symptoms to those
seen in animals, particularly primates, suggests that the
toxic effects observed in animals must be considered potentially
accurate models for humans. The persistence of symptoms
in Yusho patients is a particular source of concern. The
major uncertainty regarding toxicity in Yusho patients rests
with the unknown effects of the PCDFs present in unusually
high concentrations in Yusho oils.
Early reports of toxic effects of occupational PCB
exposure are not easily interpreted because a mixture of
compounds including chloronaphthalenes was present. A fatal
case resulted from exposure to a mixture of 90 percent chloro-
naphthalenes and 10 percent PCBs (Drinker, et al. 1937).
The subject developed chloracne, followed by jaundice and
abdominal pain, and was found to have cirrhosis of the liver
at autopsy.
Many studies of occupational exposure have shown varying
degrees of toxicity under different conditions. The following
discussion will highlight studies which indicate the types
of toxic reactions commonly observed in occupational exposures
and the levels of sensitivity in different situations. A
detailed review of occupational exposure to PCBs has recently
been prepared (Natl. Inst. Occup. Safety Health, 1977).
Elkins (1959) found that average PCB concentrations
in the workroom air of several plants in Massachusetts ranged
from 0.1 to 5.8 mg/m while peak concentrations were between
0.2 and 10.5 mg/m . No immediate toxic effects were seen;
C-59
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however, exposure to 10 mg/m was said to be unbearably
irritating. Three cases of severe chloracne were reported
in a work environment in which PCS air levels were found
to be between 5.2 and 6.8 mg/m . The workers developing
chloracne had been exposed for 2 to 4 years. No alterations
in liver function or other abnormalities were found (Puccinelli,
1954).
An analysis of the health effects of PCBs on eight
laboratory workers involved in testing dielectric fluids
was made by Levy, et al. (1977). The workers, all males
25 to 49 years of age, had been employed 2.5 to 18 years.
Breathing zone, point source, and general work area samples
were collected on three occasions. The ranges observed
were: breathing zone, 0.014 to 0.073 mg/m ; point source
(near an oven), 0.042 to 0.264 mg/m ; and room area, 0.013
to 0.15 mg/m . Blood PCB concentrations were 36 to 286
ppb which is substantially above the range in several studies
of general populations (Finklea, et al. 1972). Workers
complained of dry sore throat (6/8), skin rash (3/8), gastro-
intestinal disturbances (3/8), eye irritation and headache
(2/8). Examination disclosed one patient with skin rash,
two with nasal irritation, one showing rales, and four with
high blood pressure, but no abnormalities in liver function.
Toxic effects from a low level exposure were reported
by Meigs, et al. (1954) . A leaking heat exchanger in a
chemical plant discharged PCB vapors. No employees worked
routinely at the point of leakage but breathing zone levels
in work areas were found to be 0.1 mg/m . The period of
exposure was 19 months. Seven of 14 exposed workers developed
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mild to moderate chloracne after exposure durations of 5 -
to 14 months. Liver function tests showed normal serum
bilirubins, 24- and 48-hour cephalin flocculations, thymol
turbidities and serum alkaline phosphatase activities in
six of the seven workers, but borderline increases in cephalin
flocculation and thymol turbidity in the seventh. After
thirteen months, the thymol turbidity but not the cephalin
i
flocculation had improved.
A study of PCB exposure in six Japanese industrial
plants has been reported (Hasagawa, et al. 1972; Kara, et
al. 1974, 1975). Although the original publications are
in Japanese, a detailed description in English is available
(Natl. Inst. Occup. Safety Health, 1977). PCBs were manu-
factured in one plant, used in manufacturing capacitors
in four plants, and had been used in a fifth plant until
one month before the study began. The sixth plant used
biphenyls, not PCBs. PCB concentrations in air as both
vapor and particulates were determined. The lowest levels
3 3
in one plant were 13 to 15 pg/m vapor and 4 >ug/m particulate
while the highest levels in a single plant were 95 to 965
/ag/m vapor, 73 to 650 >ug/m particulate. Except in the
instance of a spill, vapor concentrations always exceeded
particulate concentrations. Blood PCB levels in 99 workers
were found to average 370 ppb as compared to values in 20
controls averaging 20 ppb. No correlation between duration
of exposure and blood level could be found in data from
three of the plants. Dermal effects found were chromoderm-
atosis of the dorsal joints of the hands and fingers and
of the nail bed, and acneform exanthema. Dermal effects
C-61
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seemed unrelated to blood levels, suggesting that they resulted
directly from skin contact. Changes in fat metabolism and
mild disturbances in liver function were found. The con-
sequences of termination of PCS exposure were examined by
following 38 current and 80 former workers from 1972 to
1975 who were from the plant which had discontinued PCB
use. During the period of PCB exposure, 17 capacitor immersion
process workers had blood levels of 7 to 300 ppb, which
were closely related to years of exposure. One year after
cessation of exposure, blood PCB levels decreased but not
uniformly. The average decrease was about 75 percent of
the original value. The blood half-lives of PCBs were determined
and found to be related to the number of years of exposure.
For 1 year of exposure, T*s = 3 months, while for 10 to 15
years exposure, T^s = 30 months. The investigators concluded
that blood served only as a PCB carrier while fat served
as the depot tissue. Many of the employees complained of
blackheads, acne, and skin irritation while working with
PCBs; however, these conditions cleared markedly after exposure
ceased. Serum triglyceride levels in workers were elevated
in correlation with blood PCB levels.
A study in Australia by Ouw, et al. (1976) examined
two groups of workers with different levels of exposure
in a capacitor manufacturing facility. One group (inside)
worked in an impregnation process where exposure to heated
(70 C) Aroclor 1242 occured. The second group (outside)
assembled cool Aroclor-dipped components in a location separate
from the first group. The entire group had an average blood
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PCB level near 400 ppb. The distribution of individual
Aroclor components differed between the groups with the
outside workers being low in early eluting (on gas chromatO'
graphy) fractions but elevated in late eluting fractions
relative to the inside group. No abnormalities in liver
function were observed but skin irritation and eczematous
rashes were observed. One worker had chloracne but no systemJLc
effects. The severity of dermal effects was not clearly
correlated to blood PCB level. Breathing zone air concentra-
tions in the impregnation room varied from 2.22 to 0.32
mg/m . To bring conditions within government guidelines,
improved exhaust ventilation was installed and workers were
encouraged to wear impervious gloves to reduce skin absorption.
These actions reduced atmospheric PCB levels to 0.75 to
q
0.08 mg/m . After two months, new blood samples were taken
which indicated that a slight increase in blood levels had
occurred. Failure to wear gloves was the reason cited for
the failure to improve blood levels.
A recent study of liver function in Aroclor 1016-exposed
workers illuminates the sensitivity of the liver to exposure
(Alvares, et al, 1977) . Antipyrene clearance was determined
in five workers who had been occupationally exposed to PCBs
for at least four years and Aroclor 1016 for at least two
years. None of the workers showed any manifestations of
PCB toxicity. When compared to five controls matched for
sex, age, and smoking and drinking habits, the antipyrene
half-life was about 2/3 of the control level (10.8 + 0.7
experimental vs. 15.6 + 1.0 control). The increased rate
of antipyrene clearance was taken to be an indication of
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higher levels of metabolic enzymes in the livers of the
exposed workers.
Data from this limited review of occupational studies
indicate that symptoms much like those seen after PCB inges-
tion can occur after atmospheric or dermal exposure. Air
PCB concentrations as low as 0.1 mg/m can produce toxic
effects (Meigs, 1954) and exposure to levels producing no
overt toxicity can affect liver function (Alvares, et al.
1977) . Recovery after termination of exposure occurs but
is slow and depends upon the amount of PCBs stored in adipose
tissue (Natl. Inst. Occup. Safety Health, 1977).
Synergism and/or Antagonism
It appears that the synergistic antagonistic effects
of PCBs result from their ability to induce mixed function
oxidases in liver and other tissues, although the effects
of the accelerated metabolism of drugs, such as phenobarbitol
or hormones, such as ketosteroids and thyroxin, have been
discussed above. The consequences of the PCB induced metabo-
lism of carcinogenic agents such as benzene hexachloride
or aflatoxin will be discussed below in the section on carcin-
ogenicity.
Teratogenicity
The reproductive effects of PCBs in animals and man
have been discussed above. It is clear that PCBs readily
cross the placental barrier and accumulate in fetal tissues.
Primate infants exposed to PCBs _iri utero are typically retard-
ed in growth during gestation (Barsotti and Allen, 1975)
and reproductive failures (abortions, stillbirths) are common
(Linder, et al. 1974). Live born animal and human infants
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often display symptoms of toxicity common for the species
(Kuratsune, et al. 1976; Linder, et al. 1974). However,
indications of structural malformations or genetic changes
have been rare. Villeneuve, et al. (1971b) noted assymetric
skull formation in two rabbit fetuses exposed to high levels
of Aroclor 1254 rn utero. A written communication by F.L.
Earle (as cited in Natl. Inst. Occup. Safety Health, 1977)
reported unspecified terata in canine pups born to females
exposed to 48 or 200 ppm but not 20 ppm dietary equivalent,
and in piglets from sows fed the equivalent of 50 ppm.
No additional information was given.
Mutagenicity
The mutagenicity of different PCB preparations has
been evaluated in several test systems. The single isomer
4-chlorobiphenyl was found to be highly mutagenic in Salmonella
typhimurium strain TA1538 after liver microsomal enzyme
activation (Wyndham, et al. 1976) . The products formed
under these activation conditions were 4 chloro-4'-biphenylol
and 4'chloro- 3, 4 biphenyldiol, which, as previously discussed,
are indicative of arene oxide formation (Safe, et al. 1975).
In the same study, Aroclor 1221 was less mutagenic while
Aroclor 1254, 1268 and 2, 5, 21, 5' tetrachlorobiphenyl
were essentially inactive. Mutagenic activity decreased
with increasing chlorination.
Recent attempts to repeat the experiment with different
cultures of the same tester strain have not detected any
mutagenic activity (S. Safe, personal communication).
Also 4-chlorobiphenyl was toxic but not mutagenic to
S. typhimurium TA 1538 with or without activation by Aroclor
C-65
-------
1254 (S. Rinkus, personal communication). 4-chlorobiphenyl
has been shown to induce unscheduled DNA synthesis, an indica-
tion of DNA repair, in Chinese hamster ovary cells (S. Safe,
personal communication).
The Japanese Ministry of Health and Welfare supported
mutagenicity screening program investigated Kanechlors 300
and 500 (Odashima, 1976). Both compounds were negative
in the Salmonella system but Kanechlor 300 was listed as
positive in a bacterial DNA repair assay and a cytogenetic
analysis with Yoshida sarcoma cells. Kanechlor 500 was
positive in a mouse bone marrow cell cytogenetic analysis.
Heddle and Bruce (1977) reported Aroclor 1254 as negative
in S. typhemurium, the micronucleus test and a sperm morphology
assay. Aroclor 1254 administered to rats at 50 mg/kg/day
for seven days produced no chromosomal abnormalities in
sperm (Dikshith, et al. 1975) .
The effects of Aroclor 1254 and 1242 on bone marrow
cells were evaluated in Osborn-Mendel rats (Green, et al.
1975a). Animals in groups of eight were given single doses
of Aroclor 1242 at 1250, 2000, or 5000 mg/kg or multiple
doses of 500 mg/kg/day for four days. Aroclor 1254 was
given for five days at 75, 150, or 300 mg/kg/day. Aroclor
1242 was more toxic than 1254. Mitotic indices were not
reduced by Aroclor 1242 treatment and no increase in chromoso-
mal abnormalities was observed. Aroclor 1254 reduced the
mitotic index of bone marrow cells at 150 and 300 mg/kg/day
but not at the low dose. Again, no increase in chromosomal
abnormalities was seen. Cytogenetic abnormalities were
found in spermatogonial cells of animals treated at 5000
C-66
-------
mg/kg or 500 mg/kg/day Aroclor 1242 but not in statistically
significant numbers.
A dominant lethal test with Aroclor 1242 and 1254 was
also performed in Osborne-Mendel rats (Green, et al. 1975b).
Aroclor 1242 was given in single doses of 625, 1250, or
2500 mg/kg or five doses of 125 or 250 mg/kg/day. Aroclor
1254 was given in five doses of 75, 150, or 300 mg/kg/day.
Treated males were bred to untreated females for the following
10 to 11 weeks. No significant effect of treatment was
observed on embryo implantation or lethality with any treatment,
In summary, the only marked genetic effect observed
at any level was with the single isomer 4-chlorobiphenyl.
Kanechlor 300 and 500 produced cytogenetic effects in different
systems but Aroclor 1242 and 1254 did not. Despite the
apparent weak mutagenicity of most PCBs in the systems used,
the fact that most animals can metabolize many PCB isomers
through an arene oxide intermediate indicates that the muta-
genic potential of PCBs should not be casually dismissed.
Carcinogenicity
The carcinogenic effects of PCBs have been evaluated
in several animal studies. The first evidence of carcinogenic
potential in PCBs was reported by Nagasaki, et al. (1972)
and in more detail by Ito, et al. (1973). Male dd mice
were given Kanachlors 500, 400, and 300 mixed in standard-
diets at 500, 250, and 100 ppm for 32 weeks. Of 12 mice
surviving in the group fed 500 ppm Kanachlor 500, 7 (58.3
percent) had grossly observable nodular hyperplasia with
microscopically observable hepatomas in 5 (41.7 percent).
No tumors were observed in the groups treated with lower
C-67
-------
doses of Kanechlor 500, in any dose of the other Kanechlors,
or in the six control animals. Kimbrough and Linder (1974)
treated Bald/cJ mice with Arochlor 1254. Mice were exposed
to 300 ppm in the diet for 6 or 11 months. The mice exposed
for six months were fed control diets for the remaining
five months, and all the animals were killed and examined
at the same time. All the animals surviving 11 months exposure
had enlarged livers and adenofibrosis, while 9/22 (41 percent)
were observed to have hepatomas. Of the 24 mice surviving
six months exposure, most showed some changes ifi liver cell
morphology, and a diffuse interstitial fibrosis was observed
in about 2/3 of them. One hepatoma (0.3 cm diam.) was observed.
The details of the mouse experiments are summarized in Table
10. Kimbrough and Linder (1974) reported subcutaneous abcess
formation in some mice and one sweat gland, adenoma. Neither
Ito, et al. (1973) nor Nagaski, et al. (1972) commented
on any pathology other than in the liver.
Studies with rats have been reported by Kimura and
Baba (1973), Kimbrough, et al. (1972, 1975), and Ito, et
al. (1974) . Kimura and Baba (1973) examined the effects
of Kanechlor 400 on the livers of Donryu strain rats. Ten
male and ten female animals were exposed, in a complex protocol,
to amounts of Kanechlor 400 starting at 38.7 ppm in food
and increasing to 616 ppm as the animals increased in weight.
Total amounts ingested varied from 450 to 1500 mg over exposure
periods of 159 to 560 days. Five control animals of each
sex were used. Fatty degeneration was observed in the livers
of all experimental animals and two females in the control
group. Adenomatous nodules were observed in all of the
c-68
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TABLE 10
Evidence for Carcinogenic Effects of PCB's in Mice.
n
i
vo
Mouse
Strain
dd
(Ito, et
al. 1973;
& Nagasaki
et al. 1972)
Balb/cJ
(Kimbrough
& Linder,
1974)
No.
Sex Treated
M 12
12
12
6
M 50
50
100
No.
Sur-
viving
12
12
12
6
22
24
58
PCB
Source
Kanechlor 500
n
n
Kanechlor 400
n
a
Kanechlor 300
Control
Aroclor 1254
M
n
Dietary
Level
ppm
500
250
100
500
250
100
500
250
100
300
300
Average Exposure
Daily Dose Time
mg/kg/day (Days)
82. 5a 224
41. 3a
16. 5a
82.5
41.3
16.5
82.5
41.3
16.5
49.8 330
49. 8b 180C
Liver Nodules
Adeno- Neoplastic
fibrosis Nodules Hepatoma
7/12
0/12
0/12
0/12
0/12
0/12
0/12
0/12
0/6
22/22 - 9/22
0/24 - 1/24
0/58 - 0/58
Hepatocellular
Carcinoma
5/12
0/12
0/12
0/12
0/12
0/12
0/12
0/12
0/6
<
Calculated usi.ng food consumption data from Kimbrough and Linder (1974) for Balb/cJ mice which indicates an
^average of 165 g/kg/day.
cNot directly, but assumed to be similar to group exposed 330 d.
Maintained on control diet for remaining 150 days of experiment.
-------
females which had a cumulative intake of more than 1200
mg Kanechlor 400. Nodules were seen in none of the males.
A number of histopathological findings were noted in spleen,
lung, adrenal cortex, and brain but no neoplastic changes
outside the liver were mentioned.
Ito, et al. (1974) examined the effects of Kanechlors
500, 400, and 300 on male Wistar rats. Animals were exposed
to dietary levels of 1000, 500, and 100 ppm of each preparation
for 27 to 52 weeks, then killed and examined for pathological
changes. No hepatocellular carcinoma was observed, but
cholangiofibrosis (adenofibrosis) was seen at the highest
dose of all three agents (Table 11). Nodular hyperplasia
was observed in animals treated with all three agents.
The highest incidence was observed with Kanechlor 500.
No significant changes were observed in organs other than
the liver.
Kimbrough, et al. (1975) exposed Sherman strain rats
to Aroclor 1260 at dietary levels of 100 ppm for 21 months.
t
Hepatocellular carcinomas were observed in 26/184 experimental
animals but in only one of the controls (1/173). Tumors
were observed in several other tissues of both experimental
and control groups, but they were of low incidence and fre-
quencies were similar in both groups. In an earlier study/
Kimbrough, et al. (1972) fed Aroclor 1254 and 1260 to male
and female rats for eight months. Adenofibrosis was observed
in animals fed 100 and 500 ppm Aroclor 1254 with the highest
incidence in females. Aroclor 1260 was associated with
a much lower incidence of adenofibrosis even in animals
fed 1000 ppm. A single bladder tumor was observed in a
C-70
-------
TABLE 11
Evidence for Carcinogenic Effects of PCB's in Rats.
No.
Strain Sex Treated
Donryue M 10
(Kimura
and Baba, F 10
1973)
M 5
F 5
Wistar M *
(Ito, et
al. 1974)
No.
Sur-
viving
10
10
5
5
13
16
25
10
8
16
15
19
22
18
Dietary
PCB Level
Source ppm
Kanechlor 400
Kanechlor 400
None
'None
Kanechlor 500
N
Kanechlor 400
n
n
Kanechlor 300
n
n
None
38.5-16
38.5-16
-
1000
500
100
1000
500
100
1000
500
100
0
Average Exposure
Daily Dose Time
mg/kg/day (Days)
13. 5C 339a
17. 5d 425b
-
49. Oe 378
24.5
4.9
49.0
24.5
4.9
49.0
24.5
4.9
_
Liver Nodules
Adeno-
fibrosis
-
-
4/13
0/16
0/25
2/10
0/8
0/16
2/15
0/19
0/22
0/18
Neoplastic Hepatocellular
Nodules Carcinoma
0/10
6/10
-
5/13
5/16
3/25
3/10
0/8
2/16
0/15
0/19
1/22
0/18
-
-
;
-
-
-
-
-
-
-
_
-------
Table 11 (Cont.)
Proliferative Changes
No.
No. Sur- PCB
Strain Sex Treated viving Source
Fisher M 25 24 Aroclor 1254
344 rat
(NCI, 1978) 24
24
24
? F 25 23
24
22
24
Dietary
Level
ppm
0
25
50
100
0
25
50
100
Average
Daily Dose
nig/kg/day
0
1.38e
2.75e
5.5e
0
1.38e
2.75e
5.5e
Exposure
Time
(Days)
_
735
735
735
-
735
735
735
Nodular
Hyperplasia
0/24
5/24
8/24
12/24
0/23
6/24
9/22
17/24
Hepatocellular
Carcinoma
and Adenoma
0/24
0/24
1/24
3/24
0/23
1/249
1/22
2/24
Combined
Hematopoietic
and Liver
5/24
2/24
9/24
12/24
4/23
13/24
8/22
9/24
-------
Table 11 (Cont.)
n
i
-o
u*
Strain Sex
Sherman F
(Kimbrough,
et al. F
1975)
Sherman M
(Kimbrough,
et al. F
1972)
M
F
No.
Treated
200
200
10
10
10
10
10
10
10
10
No.
Sur-
viving
184
174
10
10
8
2
10
10
10
9
Dietary
PCB Level
Source ppm
Aroclor 1260 100
None
Aroclor 1260 1000
- 100
500
1000
Aroclor 1254 100
500
100
500
Average Exposure
Daily Dose Time
mg/kg/day (Days)
4.9f 630
630
71.4 240
7.2
38.2
72.4
6.8
36.4
7.5
37.6
Liver Nodules
Adeno- Neoplastic Hepatocellular
fibrosis Nodules
144/184
0/173
2/10
1/10
1/9
4/7
1/10
10/10
7/10
9/9
Carcinoma
26/184
1/173
•
„
-
-
-
-
-
-
-
grange 159-530
"range 244-560
grange of cumulative intake 450-1800 mg
range of cumulative intake 700-1500 mg
Data not provided. Calculated from Kimbrough, et al. 1975, in which Sherman rats showed similar weight gain over the same
,experimental period.
Time weighted average calculated from Figure 2 in Kimbrough, et al. 1975.
^Reported as undifferentiated carcinoma of the liver, metastatic.
*290 animals total in 10 groups
-------
treated animal but was probably not the result of PCS exposure
(Kimbrough, et al. 1975) . The details of the experiments
with rats are summarized in Table 11.
A report dated November, 1971 described a study made
by Industrial Bio-test Laboratories Inc. A brief summary
of the report was presented at the National Conference on
Polychlorinated Biphenyls (1976) and a more detailed analysis
presented by the U.S. EPA (1976a) . One thousand Charles
river rats were divided into ten treatment groups. Fifty
male and 50 female rats served as a common control group.
Each of nine treated groups contained 50 animals of each
sex. Groups were fed 1, 10 and 100 ppm of Aroclors 1242,
1254, and 1260 respectively. Treatment was initiated with
four to six week old animals and continued for a total of
24 months. Five animals of each sex were sacrificed at
3, 6, and 12 months leaving 35 animals in each group at
the beginning of the second year. In addition, mortality
was high, leaving only 6 to 21 animals remaining in each
treatment/sex subgroup by the end of the experiment. As
seen in the previously described studies, the principal
effects were observed in the liver. Vacuolar changes and
hyperplasia were the major abnormalities originally noted
in the treated animals. In addition chromophobe adenomas
of the pituitary were found in 8 of 9 treated groups but
not in the controls. In 1975 the original liver slides
were re-evaluated with rather different results. The combined
results for animals treated with 100 ppm of all three Aroclors
included 11 hepatomas, 5 cholangiohepatomas, and 28 nodular
hyperplasias. No hepatocellular carcinomas were observed.
C-74
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Recently, a bioassay for the possible carcinogenesis
of Aroclor 1254 has been conducted by the National Cancer
Institute (1978). In this study, 24 Fischer 344 rats of
each sex were orally administered Aroclor 1254 at 25, 50,
or 100 ppm for 104 to 105 weeks. Matched controls consisted
of 24 untreated rats of each sex. Mortality among the treated
males was significantly higher than among the controls and
related to dose (P< 0.001) but was not different among the
females (P>0.05). Interstitial-cell tumors of the testes
in males and leukemias of either granulocytic or lymphocytic
type were observed frequently in both control and treated
animals. Tumors were observed in several other tissues
but their presence did not correlate with treatment. Proli-
ferative lesions of the liver were common in treated animals
but were not found in coatrols. The types and frequencies
of lesions are detailed in Table 11. They included nodular
hyperplasia in all treated groups increasing in frequency
with dose, adenomas (one male, three females) and hepatocellu-
lar carcinoma (three males, no females). In addition, adeno-
carcinomas of the stomach, jejunum or cecum of two treated
males and two treated females but no controls were observed.
Statistical analysis of the frequencies of tumors and prolife-
rative lesions indicated that the combined incidences of
leukemia and lymphoma in treated males were significant
by one test (Cochran-Armitage test for positive dose-related
trend) but not by a more stringent test (Fisher exact test).
The tumors of the liver and gastrointestinal tract were
not statistically significant; however, the occurrence of
nodular hyperplasia appeared to be related to treatment.
C-75
-------
The study concluded that Aroclor 1254 was not carcinogenic
in Fischer 344 rats; however, the high frequency of hepatocel-
lular proliferative lesions was considered to be a result
of treatment, and the carcinomas of the gastrointestinal
tract possibly associated with the treatment.
The tumors observed in rodent experiments with PCBs
were predominantly adenofibrosis (cholangiofibrosis) neoplastic
nodules and hepatocellular carcinomas. Stewart and Snell
(1957) concluded that adenofibrosis cannot be considered
to be a pre-malignant lesion, while Reuber (1968) proposed
that cholangiofibrosis might be a precursor to cholangiocar-
cinoma. Neoplastic nodules have been observed before the
appearance of carcinomas in several studies with known carcin-
ogens (Kimbrough, et al. 1975). Well-differentiated mouse
hepatomas have been shown to be potentially malignant, with
a proportion being transplantable and capable of metastasis
(Andervant and Dunn, 1952).
Several conclusions can be drawn from the results of
the rodent studies. A correlation between degree of chlorina-
tion and tumor inducing potential was observed in mice (Ito,
et al. 1973) and rats (Ito, et al. 1974) with the most highly
chlorinated preparations being most potent. However, Aroclor
1254 was more potent than Aroclor 1260 in rats (Kimbrough,
et al. 1972). Where examined, female rats were found to
be more sensitive than males (Kimura and Baba, 1973; Kimbrough,
et al. 1972). No comparisons of sex-related effects were
made in mice.
It should be noted that none of these studies was a
lifetime study. In all cases, animals were treated for
C-76
-------
fixed times then killed and examined. No lifetime studies
with PCBs were found in this survey. Such studies, if available,
might indicate more clearly the significance of the potentially
preneoplastic lesions induced by PCBs in the studies described
here.
Data on the possible carcinogenicity of PCBs in humans
are sketchy at this time. The largest group of exposed
individuals followed longitudinally are the "Yusho" patients.
By late 1973, 2 of 1291-patients had died, 9 of them with
malignant neoplasms (2 stomach cancer, 1 stomach and liver
cancer, 1 liver cancer with cirrhosis, 1 lung cancer, 1
lung tumor, 1 breast cancer, and 1 malignant lymphoma) (Urabe,
1974; Kuratsune, et al. 1976). The authors did not have
sufficient information to make a detailed epidemiological
analysis but concluded that 9/22 deaths from cancer may
represent an excess of deaths.
Two cases of malignant melanoma were reported in a
group of 31 industrial workers exposed "heavily" to Aroclor
1254 in the process of its manufacture. Based on a person-
year analysis and the use of the Third National Cancer Survey
incidence rates (Natl. Cancer Inst. 1978), 0.04 malignant
melanomas would have been expected making these data signi-
ficant at the 0.001 level. In addition, one of 41 workers
exposed to lower levels of Aroclor 1254 developed a malignant
melanoma (Bahn, et al. 1976).
Although these studies involve small numbers of indivi-
duals and provide little information about exposure or other
relevant factors, they do suggest that human exposure to
PCBs may be associated with increased risk of neoplasia.
C-77
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In addition to the carcinogenic effects observed with
PCBs, they have been shown to have a significant effect
on the carcinogenic properties of other substances found
i
in the environment. The co-carcinogenic properties of the
PCBs result from their ability to induce the mixed function
oxidases, particularly in liver, as discussed under Acute,
Sub-acute,and Chronic Toxicity. Ito, et al. (1973) observed
|hat dietary levels of 250 ppm Kanechlor 500 markedly promoted
hepatocellular carcinoma and nodular hyperplasia in mice
exposed to benzene hexachloride at levels of 100 or 250
ppm in the diet. Kanechlor 400 at 10 or 100 ppm in the
diet failed to promote cervical carcinoma or progression
toward it in mice exposed to 20 methyl cholanthrene saturated
thread implanted in the cervix and uterus (Uchiyama, et
al. 1974) . Ito, et al. (1978) observed a pronounced increase
in the incidence of preneoplastic, hyperplastic nodules
in N-2-fluorenylacetamide treated rats. The animals were
fed 1000 ppm PCB (type not specified) for eight weeks following
two weeks exposure to the carcinogen. This increase in
preneoplastic lesions over a short period was taken to be
a significant indicator of carcinogenic activity. The ability
of Aroclor 1254 to initiate (as opposed to promote) tumors
in the two-stage mouse skin system was recently examined
by DiGiovanni, et al. (1977). Aroclor 1254 proved to be
a weak initiator of papillomas when a 100 ug treatment of
skin was followed by 32 weeks of treatment with the promoter
27 tetradenanoyl-phorbol-131 acetate. When used in combina-
tion with the potent initiator dimethylbenzanthracene Aroclor
1254 slightly increased the incidence of papillomas. Aroclor
C-78
-------
1254 also failed to promote skin tumors initiated by dimethyl-
benzanthracene in the same system (100 >ug Aroclor 1254 applied
twice weekly for 30 weeks) (Berry, et al. 1978).
Kanechlor 500 promoted hepatocellular carcinoma initiated
by diethylnitrosamine (DENA) in male Wistar rats (Nishizumi,
1976). Promotion was observed when PCB treatment was begun
one week following the end of DENA treatment. The number
of tumors was significantly higher in rats treated with
DENA and PCB than DENA alone or DENA and phenobarbital although
a promoting effect was observed with the latter drug as well.
Hepatocarcinogenesis initiated by 3'-methyl - 4-dimethyl-
aminoazobenzene (3'-Me-DAB) in female Donryu strain rats
was promoted by oral administration of PCBs following initia-
tion. Tumor incidences in animals treated with 3'-Me-DAB
4- PCB, 3'-Me-DAB alone, or PCB alone were 64 percent, 13
percent, and 0 percent, respectively. PCB treatment preceding
or simultaneous with 3'-Me-DAB treatment did not produce
tumors (Kimura, et al. 1976).
By contrast to the hepatic co-carcinogenic effects
of PCBs observed by Kimura, et al. (1976) , Nishizumi (1976) ,
and Ito, et al. (1973; 1978), other investigators have observed
an inhibition of tumor formation or growth in the presence
of PCBs. Makiura, et al. (1974) fed male Sprague Dawley
rats 3'-Me-DAB, 2FAA, or DEN or pairwise combinations of
them for 20 weeks followed by 4 weeks on a stock diet.
Incidence of hepatocellular carcinoma ranged from 65.2 to
92.3 percent, and nodular hyperplasia reached 100 percent
in animals fed pairs of carcinogens. The addition of 50
ppm Kanechlor 500 to the diet resulted in a large decrease
C-79
-------
in the tumor incidence and liver weight as compared to carcin-
ogen treatment without PCBs. PCBs alone induced no tumors
or hyperplastic nodules but did result in an increased liver
weight. The principal difference between this study and
those of Ito, et al. (1978), Nishizumi (1976), and Kimura,
et al. (1976) using the same chemicals is that PCS exposure
was delayed until after the initiating treatment in the
latter studies. This suggests that the induction of mixed
function oxidases by PCS at the time of carcinogen treatment
results primarily in the inactivation of the chemicals and
that the promoting effects observed with sequential exposure
result from some other mechanism. The co-carcinogenesis
of PCBs with simultaneous exposure to BHC may reflect a
difference in the liver metabolism of this compound.
In rainbow trout (Salmo gairdnerii) 100 ppm Aroclor
1254 added to the diet reduced the size and frequencies
of liver tumors induced by 6 ppm aflatoxin B, after a one
year exposure (Hendricks, et al. 1977).
In addition to the inhibition of tumor induction by
some chemicals, PCBs were also shown to inhibit the growth
of experimental tumors in rats. Sprague-Dawley rats were
inocculated with Walker 256 Carcinosarcoma cells and the
effects of PCBs determined. Both dietary (Kerkvliet and
Kimeldorf, 1977a) and injected (Kervliet and Kimeldorf,
1977b) Aroclor 1254 reduced the size of solid tumors and
increased animal lifespan. Total dietary PCB intake of
1100 to 2000 mg/kg over a 40-day period reduced tumor weight
to 60 to 40 percent of control in both male and female rats.
Aroclor 1254 injected i.p. reduced the efficiency of tumor
C-80
-------
takes when 10 tumor cells were injected from 81.3 in control
to 50.0 percent in animals receiving 200 mg/kg/day. Mean
tumor sizes were reduced and lifespans increased by PCBs
in animals inocculated with 10 tumor cells. Administration
of PCBs for five days preceding tumor inoculation or the
first five days after inoculation was more effective than
administration between days five and ten.
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CRITERION FORMULATION
Existing Guidelines and Standards
The Toxic Substances Control Act (TSCA) (P.L. 94-469)
was signed into law October 11, 1976. Provisions in section
6(e) of the law specifically regulate the manufacture/ sale/
distribution, and disposal of PCBs* Manufacture, sale,
or distribution of PCBs was restricted to sealed systems
as of October 11, 1977. Manufacture was banned as of January
1, 1979 and all processing and distribution in commerce
will cease July 1, 1979. Allowance for certain exemptions
is provided in the law. The proposed rules to implement
the terms of section 6(e) of TSCA were released June 7,
1978 (U.S. EPA, 1978b). Proposed rules on the disposal
of PCBs were released February 17, 1978 (U.S. EPA, 1978a).
The Environmental Protection Agency has proposed a water
quality criterion for the protection of fresh water and
marine life of 0.001 ug/1 (U.S. EPA, 1976b). The Food and
Drug Administration established tolerance levels in foods
in 1973 (38 FR 18096) and proposed new tolerance levels
further restricting levels in 1977 (42 FR 17487). Both
the current allowed levels and the proposed levels are presented
in Table 3.
The occupational exposure limits adopted in 1968 are
based on the recommendations of the American Conference
of Governmental Industrial Hygienists (ACGIH) (1968). They
set the time-weighted average eight hour exposure limits
to .1.0 mg/m for mixtures containing 42 percent chlorine
and .5 mg/m for mixtures containing 54 percent chlorine.
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The newly recommended standard proposed by NIOSH (1977)
is 1.0 ug/m air TWA over a 10-hour day and 40-hour work
week.
Current Levels of Exposure
Human exposure to PCBs in the United States has been
broad. Several studies of tissue and plasma levels of PCBs
have detected them in!a high percentage of randomly chosen
subjects. Yobs (1972) detected PCBs in 31.1 percent of
637 human adipose tissue. The National Human Monitoring
Program for Pesticides in fiscal years 1973 and 1974 found
PCBs in 35.1 and 40.3 percent of adipose tissues tested
(Kutz and Strassman, 1976). Table 12 indicates the distribu-
tion of PCB concentrations in the population. A study of
Canadian human adipose tissue PCB levels found 1 ppm or
more in 30 percent of 172 samples (Grant, et al. 1976).
The eastern provinces, particularly Ontario, had the highest
incidences. Average adipose tissue PCB levels were just
below 1 mg/kg (ppm) with males having slightly higher accumu-
lations than females. The same study found human breast
milk to contain about 1 mg/kg on a fat basis. PCBs were
detected in 8 of 40 samples of breast milk in Colorado at
levels between 40 and 100 ppb (whole milk). The Japanese
study described earlier found average levels in 400 milk
samples of about 30 ppb (Yakushiji, et al. 1977). PCB levels
in plasma in U.S. populations were detected in 43 percent
of 723 samples. Levels in positive samples ranged from
1.5 to 29 ppb with a mean around 2 to 3 ppb. White populations
had higher levels than black populations (Finklea, et al.
1972).
C-83
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TABLE 12
Levels of Polychlorinated Biphenyls in Human Adipose Tissue
Data Sample Percent Percent Percent Percent
source size nondetected 1 ppm 1-2 ppm 2 ppm
Yobs, 688 34.2 33.3 27.3 5.2
1972
FY 1973 1277 24.5 40.2 29,6 5.5
Survey
FY 1974 1047 9.1 50.6 35.4 4.9
Survey
From: Kutz and Strassman (1976)
C-8'4
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As discussed in the section on exposure, the median
water levels of PCBs are around 0.1 to 0.3 jug/1 in positive
samples with 0 to 20 percent of samples being positive around
the U.S. (Dennis, 1976). Average PCB intake in food was
estimated in the mid-1970's to be about 9 jug/day with fish
being the major dietary source. Ambient air concentrations
are around 100 ng/m (Kutz and Yang, 1976).
Special Groups at Risk
The preceding discussion of human exposure makes clear
the fact that a high percentage of the U.S. population has
been and is exposed to low levels of PCBs in food, water,
and air. Those groups at particular risk for PCB exposure
include industrial workers exposed in the workplace, indivi-
duals consuming large amounts of contaminated fish, such
as sport fisherman (42 FR 17487), and nursing infants who,
per kg body weight, may accumulate significant body burdens
from the levels in human breast milk. With the cessation
of manufacture of PCBs by Monsanto in 1977 and the great
decline in its use which should result from the implementation
of section 6 (e) of TSCA, industrial exposure should decline
substantially. Since many PCB-containing sealed systems
can be expected to remain in service for many years continuing
vigilance will be necessary to minimize accidental pollution
of waterways or air and to prevent further occupational
exposure.
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Basis and Derivation of Criterion
In arriving at .a.criterion for PCB levels in ambient
waters several factors must be taken into account. First,
PCBs are highly persistent in the environment and accumulate
to a high degree in food webs. As discussed in the section
Ingestion from Foods, an average bioaccumulation factor
for PCB's in all freshwater fish and shellfish of 46,000
has been determined. As a consequence, PCBs leave the envi-
ronment very slowly once they have entered it. Not only
do PCBs persist and accumulate in the environment but in
man as well. The current environmental levels are not produc-
ing obvious acute ill health in the general population.
However, several animal studies report that PCBs produce
a carcinogenic response and that they may enhance the carcino-
genic activities of other substances.
Although other adverse effects of PCB exposure could
be used as a basis for formulating a criterion, carcinogenicity
will be used for a variety of reasons. The most extensive
chronic studies with PCBs have identified carcinogenicity
as the major end point. Although no carcinogenicity studies
have been extended to more than one generation and firm
data exist only for the female rat, a credible carcinogenic
response to PCBs has been demonstrated and cannot be ignored.
Kimbrough, et al. (1972) observed an incidence of hepatocellular
carcinoma of 26/184 in treated rats compared to 1/173 in
controls. The NCI bioassay observed a similar frequency
of hepatocellulor carcinoma at a similar dose level which
was statistically not significant because the number of
animals was low. In addition, a number of non-malio..
C-86
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proliferative processes observed in liver at high frequencies
in the PCB treated animals in these studies were also observed
in both rats and mice in other studies (Ito, et al. 1974;
Kimura and Baba 1973; Kimbrough, et al. 1972; Ito, et al.
1973; Kimbrough and Linder, 1974). PCBs were classified
as carcinogenic by the International Agency for Research
on Cancer (IARC, 1974). Evidence from human populations
suggests but does not confirm an increase in cancer frequency
due to PCB exposure (Kuratsune, et al. 1976; Bahn, 1976).
Finally, a theoretical basis exists for the quantitative
extrapolation of carcinogenic effects in treated animals
to human populations. Various models, such as the one used
below, can provide quantitative risk estimates based on
animal data, and certain assumptions about the induction
of neoplasia (e.g. one-hit or multi-hit induction). No
basis exists for extrapolation with mathematical models
from animals to man for many other kinds of biological effects.
Although the criterion established below is based on
animal carcinogenicity data it should also be .noted that
other adverse effects have been observed in mammals at levels
below the dose which produces tumors in rats. The carcinogenic
effect was observed in rats consuming an average of 4.9
mg/kg/day. Dietary levels at 2.5 ppm produced adverse repro-
ductive effects in Rhesus monkeys (Allen and Barsotti, 1976).
If a food intake of 350 g/day is assumed, the PCB dose is
146 ug/kg/day in 6 kg animals. At this time no data are
available to indicate the minimal level in the diet at which
PCBs produce toxic effects in Rhesus monkeys.
C-87
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in mink, ingestion of as little as 61 mg of Aroclor
1254 over nine months or 90 mg of Aroclor over four months
resulted in sharply reduced reproduction (Aulerich and Ringer,
1977) . Assuming a weight of 1 kg for adult mink and a food
intake of 150 g/day, the PCB dose at 2 ppm was about 300
ug/kg/day which is similar to the level producing reproductive
toxicity in monkeys.
These data can be used in one approach to developing
an ambient water quality criterion. If 300 jug/kg/day is
taken as the lowest-observable-effect-level (LOEL) than
an Acceptable Daily Intake (ADI) can be calculated for a
70 kg man using an uncertainty factor of 100:
7°
Assuming that exposure to PCBs is based on the consumption
of 2 liters of drinking water, 18.7 grams (0.0187 kg) of
fish and shellfish, and a bioconcentration factor of 46,000;
then the following calculation can be made:
(2 liters) X + (0.0187 x 46,000) = 210 jug
X = .244 jug/1
or 244 ng/1)
As will be seen later, the carcinogenicity criterion
methodology gives a lower and presumably more cautionary
level.
An assessment of carcinogenic risk will be made by
extrapolation from animal data using a linear (non-threshold)
model. The model used takes into account the bioaccumulation
of PCBs in fish and shellfish. It is assumed that an average
of 2 liters/day of water are consumed along with 18.7 g
of fish taken from that water source. Exposures from ~>t-her
C-88
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food sources, air or occupational exposure are not included
in the criterion level derived by this model.
Among the studies reviewed by this document, only one
appears suitable for use in the cancer risk assessment.
None of the mouse studies involved feeding for most or all
of a lifetime and are therefore unsuitable. Of the rat
studies, the only one involving long term exposure and adequate
numbers of animals is the study of Sherman rats by Kimbrough,
et al. (1975).
This study has some drawbacks in that it lacks any
evidence of a dose-response (due to the use of only one
dose level), it tests only one sex of the species, and only
one commercial mixture of PCBs was tested. Yet the experi-
mental design is a good one in many ways: the treatment
was given over a good proportion of the lifespan; there
was an appropriate route (food) and distribution of exposure
(uniform dose over time); the authors provided good documenta-
tion of the actual intake dose; a sufficiently large number
of experimental and control animals were used to detect
a statistically significant increase in tumors; and there
was a thorough and well documented description of the pathology
(hepatocellular carcinoma). The NCI study (1978) was the
only other study involving a long-term exposure and was
suggestive of a carcinogenic effect; however, the lack of
an adequate number of animals renders it unsuitable as a
study upon which to base an estimate of carcinogenic risk.
Under the Consent Decree in NRDC vs. Train, criteria
are to state "recommended maximum permissible concentrations
(including where appropriate, zero) consistent with the
C-89
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protection of aquatic organims, human health, and recreational
activities." PCBs are suspected of being human carcinogens.
Because there is no recognized safe concentration for a
human carcinogen, the recommended concentration of PCBs
in water for maximum protection of human health is zero.
Because attaining a zero concentration level may be
infeasible in some cases and in order to assist the Agency
and States in the possible future development of water quality
regulations, the concentration of PCBs corresponding to
several incremental lifetime cancer risk levels have been
estimated. A cancer risk level provides an estimate of
the additional incidence of cancer that may be expected
in an exposed population. A risk of 10 for example, in-
dicates a probability of one additional case of cancer for
every 100,000 people exposed, a risk of 10 indicates one
additional case of cancer for every million people exposed,
and so forth.
In the Federal Register notice of availability of draft
ambient water quality criteria, EPA stated that it is con-
sidering setting criteria at an interim target risk level
of 10~5, 10~6 or 10~7 as shown in the table below.
Exposure Assumptions Risk Levels and Corresponding Criteria (1)
(per day) _7 _fi -
0_ 1£ ' 10 ° 1£ D
2 liters of drinking water 0 0.0026 ng/1 0.026 ng/1 0.26 ng/1
and consumption of 18.7
grams fish and shellfish. (2) .
Consumption of fish and 0 0.0026 ng/1 0.026 ng/1 0.26 ng/1
shellfish only.
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(1) Calculated by applying a modified "one-hit" extrapolation
model described in FR 15926, 1979. Appropriate bioassay
data used in the calculation are presented in Appendix
I. Since the extrapolation model is linear at low
doses, the additional lifetime risk is directly propor-
tional to the water concentration. Therefore, water
concentrations corresponding to other risk levels can
be derived by multiplying of dividing one of the risk
levels and corresponding water concentrations shown
in the table by factors such as 10, 100, 1,000, and
so forth.
(2) Approximately 99.8 percent of the PCS exposure results
from the consumption of aquatic organisms which exhibit
an average bioconcentration potential of 46,000 fold.
The remaining 0.2 percent of PCB exposure results from
drinking water.
Concentration levels were derived assuming a lifetime
exposure to various amounts of PCBs, (1) occurring from
the consumption of both drinking water and aquatic life
grown in waters containing the corresponding PCB's concentra-
tions and, (2) occurring solely from consumption of aquatic
life grown in the waters containing the corresponding PCB
concentrations, although total exposure information for
PCBs is discussed and an estimate of the contributions from
other sources of exposure can be made, this data will not
be factored into ambient water quality criteria formulation
until additional analysis can be made. The criteria presented,
therefore, assume an incremental risk from ambient water
exposure only.
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These criteria are exceedingly low. Although sharp
restriction of open PCB use in 1970 resulted in notable
declines in water PCB levels in the next several years (Dennis,
1976), the residual levels remaining are still two to three
orders of magnitude above the extrapolated level indicated
by the model. The major source of PCBs in water today is
probably not new effluents from industrial or domestic sources,
but the PCB containing sludges underlying waterways which
typically contain 100 to 1000 fold higher concentrations
than the water itself (Dennis, 1976). Efforts to reduce
water levels significantly by eliminating current pollution
sources will probably have little effect on average water
PCB concentrations.
The very low limits suggested by this risk estimate
are due in large part to the very large bioaccumulation
factor in fish (46,000). This figure is an average for
a wide variety of saltwater and freshwater organisms (see
section on Ingestion from foods).
As possible strategies to reduce human exposures to
PCBs are considered, the relative contributions of ingested
water and fish should be kept in mind. At the assumed consump-
tion rate of 2 1 of drinking water and 18.7 g of fish/day,
over 99 percent of the dietary PCBs will be obtained from
fish. Strategies which focus separately on the reduction
of PCB levels in water and fish for human consumption might
be more practical and productive than a single standard
for water which takes bioaccumulation in fish into account.
A final comment about the risk level derived from this
study is that it is based on animal data which are st.~t.is-
C-92
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tically weak. The weight of evidence indicates that PCBs
are carcinogenic in rodents. However, the carcinogenic
activities of these compounds are not great. An acceptable
noncarcinogenic level could be established with greater
certainty if better quantitative data on carcinogenicity
were available. Studies with larger numbers of animals
designed to measure relatively small effects are needed.
Also, the rat appears to be much less sensitive to the acute
and subacute effects of PCBs than man or non-human primates.
Further investigation of the effects of PCBs in Rhesus monkeys,
particularly with reference to the gastric lesions produced,
would be useful.
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APPENDIX I
Summary and Conclusions Regarding the Carcinogenicity of
Polychlorinated Biphenyls*
Polychlorinated biphenyls (PCBs) are prepared by the
chlorination of biphenyl and are complex mixtures containing
isomers of chlorobiphenyls with different chlorine content.
Because of the widespread industrial use of PCBs, their
long half-life, and the documented disease-producing capability
of these compounds in several species, regulations have
been promulgated banning most of the manufacturing, processing,
and distribution of PCBs in the United States (FEDERAL REGISTER
Vol. 44, No. 106, May 31, 1979).
Human studies concerning the possible carcinogenicity
of PCBs have involved small numbers of individuals and provide
little information about exposure. Although these studies
are only marginally useful in describing the carcinogenicity
of PCBs, the incidence of malignant neoplasms in "Yusho"
patients and in industrial workers exposed to Aroclor 1254
suggests that human exposure to PCBs is associated with
an increased risk of neoplasia.
In two separate studies, PCBs have been reported to
induce hepatocellular carcinomas in both mice and rats (male
mice fed Kanechlor 500 at 500 ppm and female Sherman rats
fed Aroclor 1260 at 100 ppm).
In an NCI bioassay, Aroclor 1254 was not carcinogenic
in Fischer 344 rats, but the high frequency of hepatocellular
proliferative lesions was considered to be the result of
treatment and carcinomas of the gastrointestinal trac'- oossibly
C-94
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associated with treatment. In one other mouse study and
three other rat studies, various PCBs induced proliferative
lesions of the liver which might be indicative of carcino-
genicity. The most commonly seen lesions were adenofibrosis
(cholangiofibrosis) and neoplastic nodules.
A correlation between degree of chlorination and tumor
inducing potential was observed in both mouse and rat species.
The most highly chlorinated preparations were also the most
potent tumor inducers with the exception of Aroclor 1254
which was more potent than Aroclor 1260 in one rat study.
Where examined, female rats were found to be more sensitive
than males. No comparisons of sex related effects were
made in mice.
PCBs have been reported to be co-carcinogens, initiators,
and promotors in both mouse and rat species.
The mutagenicity of different PCB preparations has
been evaluated in several test systems with conflicting
results. In one study, the single isomer 4-chlorobiphenyl
was reported to be highly mutagenic in Salmonella typhimurium
strain TA 1538 after liver microsomal activation, while
Aroclor 1221 was reported to be less mutagenic and Aroclors
1254, 1268, and 2,5,2',5'-tetrachlorobiphenyl were inactive.
The fact that mutagenic activity decreased with increasing
chlorination is consistent with the characteristic insensitivity
of the ames test to chlorinated hydrocarbons. In other
test systems, Kanechlor 300 inhibited bacterial DNA repair
deficient cells and induced cytogenetic abnormalities in
Yoshida sarcoma cells. Kanechlor 500 tested positive in
a mouse bone marrow cytogenetic analysis.
C-95
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In summary, carcinogenic responses have been induced
in mice and rats. These results, together with positive
mutagenic responses, and suggestive epidemiologic evidence,
constitute substantial evidence that PCBs are likely to
be human carcinogens.
The water quality criterion for PCBs is based on the
Kimbrough, et al. (1975) study on the induction of hepatocellular
carcinomas and neoplastic nodules in female Sherman strain
rats fed 100 ppm Aroclor 1260. It is concluded that the
water concentration of PCBs should be less than 0.26 ng/1
(~'0.2 ng/1) in order to keep the lifetime cancer risk below
io-5.
*This summary has been prepared and approved by the Carcinogens
Assessment Group of EPA on June 15, 1979.
C-96
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Summary of Pertinent Data
The water quality criterion for PCBs is derived from
the hepatocellular carcinoma and neoplastic nodule response
of Sherman strain female rats fed 100 ppm Aroclor 1260 (Kimbrough,
et al., 1975). A time-weighted average dose of 88.4 ppm
was administered for approximately 21.5 months and the animals
were observed for an additional six weeks before terminal
sacrifice. The incidence of hepatocellular carcinoma and
neoplastic nodules was 170/184 in the treated group and
1/173 in the control group. Assuming a fish bioaccumulation
factor of 46,000, the criterion is calculated from the following
parameters:
n. = 170 d = 88.4 x 0.05 = 4.42 mg/kg/day
N£ = 184 w = 0.4 kg
n = 1 L = 730 days
N:; = 173 R = 46,000
Le = 730 days F = 0.0187 kg/day
le = 645 days
Based on these parameters, the one-hit slope BH is
3.25 (mg/kg/day)~ . The resulting water concentration of
PCBs calculated to keep the individual lifetime cancer risk
below 10~ is 0.26 nanograms per liter.
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