Cancer Risk: Specific Dose Estimate for 2,3,7,8-TCDD, Appendices A-F, External Review Draft

'United States
Environmental Protection
Agency
Office of Health arid
Environmental Assessment
Washington DC 20460
EPA/600/6-88/007Ab
June 1988
External Review Draft
Research and Development
A Cancer
Risk-Specific
Dose  Estimate for
2,3,7,8-TCDD

Appendices A
Through F
                 Review
                 Draft
                 (Do Not
                 Cite or Quote)
               NOTICE
 This document is a preliminary draft. It has not been formally
 released by EPA and should not at this stage be construed to
 represent Agency policy. It is being circulated for eomment on its
 technical accuracy and policy impticati'ons;

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(Do Not Cite or Quote)                            EPA/600/6-88/007Ab
                                                   June 1988
                                            External Review Draft
                      A Cancer
    Risk-Specific Dose  Estimate for
                  2,3,7,8-TCDD
                           NOTICE
This document is a preliminary draft. It has not been formally released by EPA and should
not at this stage be construed to represent Agency policy. It is being circulated for comment
on its technical accuracy and policy implications.
              U.S. Environmental Protection Agency
                      Washington, DC

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                                DISCLAIMER
     This document is an external draft for review purposes only and does
not constitute Agency policy.  Mention of trade names or commercial products
does not constitute endorsement or recommendation for use.

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                                                                           March  1988
                                                                           Review  Draft
                                           APPENDIX  A
                        QUANTITATIVE  IMPLICATIONS  OF  THE USE OF DIFFERENT

                        EXTRAPOLATION  PROCEDURES FOR  LOW-DOSE CANCER  RISK

                             ESTIMATES FROM EXPOSURE  TO 2,3,7,8-TCDD
                                     Steven  P.  Bayard,  Ph.D.
                                   Carcinogen Assessment Group
I
                          Office of Health and Environmental  Assessment
                                Office of Research and Development
                               U.S. Environmental  Protection  Agency

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                                    .. CONTENTS


  I.    INTRODUCTION AND DEFINITION OF TERMS.  .	1

       A.   INTRODUCTION	1

       B.   DEFINITION OF TERMS  .	6

           1.   Terms Associated with Modeling	6

           2.   Terms Associated with the  Uncertainty Factor Approach  	 8

 II.    EPA's  USE  OF THE LINEARIZED MULTISTAGE MODEL FOR  CARCINOGEN
       RISK EXTRAPOLATION AND COMPARISON  WITH OTHER MODELS  	 8
       A.   DESCRIPTION  OF  THE  MULTISTAGE  AND  LINEARIZED  MULTISTAGE
           MODELS	,	8

       B.   USE OF  THE LINEARIZED  MULTISTAGE MODEL FOR  RISK EXTRAPOLATION
           OF  2,3,7,8-TCDD:  COMPARISON OF FOUR  U.S. AGENCIES	9

       C.   ALLOMETRIC AND  BODY BURDEN CONSIDERATIONS	  .  .13

       D.   OTHER  EXTRAPOLATION MODELS	18

           1.   Longstreth  and  Hushon (1984)	  .18

           2.   Sielken  (1987).	  .  .22

       E.   TIME-TO-TUMOR ANALYSES	26


III.   TREATMENT  AS A PROMOTER	28

       A.   UNCERTAINTY FACTOR  APPROACH	28

       B.   MODELING AS  A PROMOTER UNDER  THE MOOLGAVKAR,  VE-NSON,
           AND KNUDSON TWO-STAGE MODEL	  .31


 IV.   COMPARISON OF ANIMAL PREDICTION WITH  ACTUAL HUMAN DATA. .  .	35

       A.   THIESS et al. (1982)	38

           1.   Description	38

           2.   Exposure Estimates	38

           3.   Risk Estimates	•	41

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                               CONTENTS  (continued)
 V.



VI.
B.  ZACK AND SUSKIND  (1980)



    1.  Description  .  .  .  .



    2.  Risk Estimates.  .  .



DISCUSSION AND SUMMARY.  .  .



REFERENCES	
.44



.44



.44



.49



.54
                                       m

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                    I.   INTRODUCTION AND DEFINITION  OF  TERMS

A.  INTRODUCTION
     2,3,7,8-Tetrachlorodibenzo-jD-dioxin (2,3,7,8-TCDD)  is  the most  potent
animal  carcinogen ever  tested.   It  is 50 times  more  potent  than  aflatoxin Bl on
a per mole basis and 50 million times more potent than  vinyl  chloride.   In
addition to its carcinogenic potency, 2,3,7,8-TCDD is also  the most  potent  animal
teratogen known and it  causes other reproductive and immune system effects  at
extremely low doses as  well.
     Because of these severe toxicities, many U.S. federal  and  state, as well  as
foreign, regulatory and health  agencies have proposed or implemented regulations
or advisories based on  levels of concern for 2,3,7,8-TCDD.   Among these agencies,
the U.S. Environmental  Protection Agency (EPA)  was,  to  this writer's knowledge,
the first to actually produce an (upper-limit)  estimate of cancer risk  for
2,3,7,8-TCDD exposure to humans (U.S. EPA, 1980).  This estimate was based  on  a
methodology that extrapolated from cancer responses  at  doses of 1, 10,  and  100
ng/kg-day in an animal  lifetime feeding study (Kociba et al., 1978;  Table 1) to
humans at still lower levels.  Initially, this  extrapolation  was based  on a
simple linear extrapolation from the lowest dose to  show a significant  elevated
response (10 ng/kg-day caused a statistically significant increase in  liver
tumors--hyperplastic nodules—versus control).   Shortly afterward, however, the
methodology was modified to include  all dose-response points in the extrapolation
procedure.  The new methodology was based on a multistage model  for carcino-
genesis due to  a specific time ordering of changes,  first proposed by  Armitage
and Doll (1954) but modified by Crump et al.. (1977,  1979) to include dose-
response for extrapolation  purposes.  Often called the linearized multistage
model, the Crump model  is distinguished by its approach of providing upper-limit

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estimates of risk consistent with nonthreshold low-dose linearity.   This model
is presented in section 11.A.
     Following EPA's efforts, two other U.S. agencies, the Food and Drug
Administration (FDA, 1983) and the Centers for Disease Control  (CDC)  (Kimbrough
et al., 1984) produced cancer risk estimates based on slight modifications of
EPA's methods but with similar resulting estimates.  In another minor variation,
the State of California (1984) used the Crump linearized multistage model  to
extrapolate the cancer response to humans from a mouse gavage study performed
by the National Cancer Institute (NTP, 1982).  All efforts produced results
within a factor of 10.  These are discussed in section II.B.
     Within the framework of the linearized multistage model and the Kociba et
al. rat feeding study, two other efforts appearing in the literature are note-
worthy.  First, Longstreth and Hushon (1984) applied several mathematical  non-
threshold, nonlinear model s (Logit,  Probit, Weibull, and multihit)  to the
cancer response in the Kociba et al. study, and compared the extrapolated
results with those of the linearized multistage model.  Second, Sielken  (1987)
fit the Kociba data with the multistage model (but not the Crump linearized
version applying upper limits) and also with a modified version which allowed
the input of actual observation times.  He then compared actual estimates
derived from the multistage model  with those of the EPA, which  used the  upper
limits based on the Crump version.  The Sielken paper is discussed  further in
section II.D.
     In contrast to all of the above attempts at extrapolating  from animal  data
to humans by nonthreshold models, several  U.S. nonregulatory agencies have
applied safety or uncertainty factors, not models.  The uncertainty factors of
between 100  and 1,000 are applied to doses that have shown no adverse effects
in animal  cancer or other studies, and the resulting numbers are presumed safe

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for humans.  This methodology  has  been  used by  EPA and many other agencies for
animal-to-human  extrapolations  for toxic effects other than cancer, but EPA has
never used this  methodology to estimate cancer  risk.  The estimates of cancer
risk provided by the uncertainty  factor approach are in the range of 150 to
1,500 times lower than the estimates  provided by EPA's use of  Crump's multistage
model.  These estimates are presented in section III.A.
     The differences in the magnitude of the estimates provided  by these two
approaches require a closer look  at the methodologies involved in each and in
the reasoning as to which is the  proper one to  use  for cancer  risk extrapo-
lation of 2,3,7,8-TCDD.  The argument focuses on the model for complete carcin-
ogens versus promoters.  Complete carcinogens  have  both  initiating and promoting
ability, and it is the mechanism  leading  to the initiating part  of carcinogenesis,
the attachment of the carcinogen  to the DNA,  that  can  be modeled on  either a
linear or multistage basis.  Those in favor of modeling  2,3,7,8-TCDD as a
complete carcinogen argue that 2,3,7,8-TCDD causes  rare  cancers  of the  hard
palate and nasal turbinates, tongue, (in male rats),  and a  rare  form of lung
cancer, and  that  such  rare tumors would be unlikely to be initiated  except by,
the  2,3,7,8-TCDD  in the experiment.  Conversely, those in favor  of the  uncertainty
factor  approach  point  to  the  strong  evidence for the promoting effects  of
2,3,7,8-TCDD in  the liver where most of the tumors are occurring.  They  argue
that promotion  is effectively  a toxic  reaction with a threshold and  that all
promoters  show  not  only thresholds but also reversibility upon cessation  of
dosing.   Treating 2,3,7,8-TCDD as a  promoter and using an uncertainty factor
approach  has a  further advantage  of  comparing  its cancer effects with its other
 toxic effects using the same  methodology.
      A third approach is  also  possible.. This  is an approach which models for
 the cancer effects of 2,3,7,8-TCDD through its known mechanism  of binding to a

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 receptor.   (This actual modeling and results are presented in section  IV.B.)
 While the basic model, the Moolgavkar-Venzon-Knudson (M-V-K)  two-stage model,
 with a  promotion phase, has been used in the literature to explain many  cancers,
 and has been found to predict well  the tumor promotion in mouse skin  (Chu  et
 al., 1987), the approach is new in that modeling for promoters has never been
 done before by regulatory agencies.
     The purpose of the presentations that follow is to compare quantitatively
 the estimates derived from each of the separate approaches and then to compare
 them with each other.  In order to do this, common terms must be introduced.
B.  DEFINITION OF TERMS
1.  Terms Associated with Modeling  (defined here as an estimation  of  the  incre-
mental cancer risk to humans arrived at by fitting a mathematical function to
animal response data)
     Maximum likelihood estimate (MLE)—The statistical  procedure by which the
parameters of the model are estimated.  The MLE has many properties, in a sta-
tistical sense, which allow it to be referred to as a "statistical  average"  or
"best" estimate.  In risk terms it might be thought of as a  term that, if the
assumed model is true, provides overestimates and underestimates of the true
risk each 50% of the time.
     Parameter—A constant in the model, associated either with the control
response, or the time or dose variable inputs.   For example,  in the Crump
linearized multistage model, the parameter associated with the linear  dose
variable is denoted as c\i and is defined as the increase in  cancer  risk asso-
ciated with an incremental increase per unit of dose.  For this reason qj is
expressed in units of reciprocal dose such as (ng/kg-day)'1.
     Upper-confidence limit (UCL) estimates—The estimates resulting from a

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statistical  procedure in which the  upper-limit  values of the parameters still
consistent with the data are  estimated.   In the linearized multistage model
(Crump), the upper-limit estimate associated  with  the linear term  is designated
q^.  In a statistical  sense it is the 95%  upper-limit estimate of the linear
term associated with the fitting of the  linearized multistage model to the
animal  data.  In making  cross-species extrapolations to humans, however, the
"95%" label  is dropped,  since the uncertainty associated with cross-species
extrapolations is considered  far greater than the  statistical uncertainty
associated with the model-fitting procedure.  Also, because the linearized
multistage model becomes linear at  low doses, the  UCL on tf[ is the same as
the UCL on the incremental  risk. This is  not true of the nonlinear models such
as the Logit, Probit, and Weibull discussed  in  section  II.  Upper-limit incre-
mental  risk estimates, however, are comparable, and the ratio of these estimates
from two different models can be expressed as the  relative potency.
     Risk specific dose (RsD)--A dose associated with a specified cancer risk.
For example, assume a linearized multistage model  is  fit to the data and the
parameter estimates are q-j_  =  3.0 x  10"3  (ng/kg-day)'1,  qi = 0 for  all i^l,
and q| = 7.5 x 10~3 (ng/kg-day)"1.   Then for  an incremental  risk of 1 in
1,000,000, the dose would be the solution  to  10~6  = l-exp(-3.0 x 10-3 d),  and
RsD = 3.3 x ID"4 ng/kg-day would be called the  risk specific dose.  Likewise,
the RsD could be defined in terms of the lower  limit  of the dose corresponding
to a risk of 10~6.  In this case aft would  be  substituted for q± and the solu-
tion would be RsD = 1.3 x 10~4 ng/kg-day.
     A ratio of two RsDs can also be thought  of as a  measure of relative potency,
but in this case the higher the RsD, the lower  the potency.  The RsD thus
becomes a common unit to discuss relative potency  between different approaches
and different types of toxicity.

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     Virtually safe dose (VSD)--A dose associated with a  very  small  risk.  The
general  reasoning in discussing RsDs and VSDs  is  identical.  The only difference
is in one's definition of a "very small  risk."
2.  Terms Associated with the Uncertainty Factor Approach
     The lowest-observed-adverse-effect-level  (LOAEL)  is  defined  as  the  lowest
dose in an experiment at which there is a statistically significant  increase
over the control group in the proportion of animals for which  adverse effects
are observed.  The no-observed-adverse-effect-level (NOAEL)  and no-observed-
effect-level (NOEL) are straightforward negations (Crockett and Crump,  1986).
The uncertainty or safety factor is an arbitrary factor applied to these levels
for the purpose of establishing concern or no-concern levels for  humans.

     II.  EPA's USE OF THE LINEARIZED MULTISTAGE MODEL FOR CARCINOGEN RISK
                 EXTRAPOLATION AND COMPARISON WITH OTHER  MODELS
A.  DESCRIPTION OF THE MULTISTAGE AND LINEARIZED MULTISTAGE MODELS
      EPA's  reasons for using the linearized multistage model for risk extrapo-
lalation, in general, are discussed in the Guidelines for Carcinogen Risk
Assessment  (U.S.  EPA, 1986).   For the 2,3,7,8-TCDD risk assessment, additional
discussions are presented in the Health Assessment Document (HAD) for Polychlor-
inated-Dibenzo-£-Dioxins (U.S.  EPA, 1985) as well as in the document on the
cancer  risk-specific dose estimate for 2,3,7,8-TCDD.  Therefore, only an abbrevi-
ated  review of its development will be presented  here.  Basically, its genesis
came  from Armitage and Doll who proposed a theory that a cancer cell was generated
from  a  series  of  several heritable mutations in a specific  order, the end result
of each change being termed a  stage.  The transition rate from one stage to the
next  was hypothesized  as being related to a probability of  occurrence.  The time

                                       8

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rate of occurrence of the ith event is a-j+b-jd-j, i=l,  ......  k, where a^ ,  b-j >  0

and d=dose.  This, along with some other assumptions, leads to a dichotomous

response probability of the form

                                k
                P(d) =l-exp-[cir  (a-j+bid)]     a-jbi £ 0
                               1 = 1             .

where a-j is the background transition rate for a progression to stage  i ,  b-j

is the incremental increase in that rate per unit of dose, c is a function  of

exposure duration, and P(d) is the probability of a tumor  by some fixed age  t

for a dose d.  This model achieved some popularity mainly  because of its  success

at predicting many of the human epithelial  cancers, and because the model now

presented the probability of a tumor as a function of dose.  In addition, the

reparameteri zation of the individual  transition rates leads to


                  P(d) = l-exp-(q0 + q-^d + q2d2 + ... + qkdk)
where P(d) is the lifetime probability of cancer at dose d.   Since qu is  the

parameter associated with the background rate, an assumption of independent

background leads to


                 Pt(d) = l-exp-(q1d + ... + qkdk)     all qn- >. 0


where Pt(d) is the incremental (often called the extra)  risk associated with

dose d.   In the linearized form of this model an upper-limit estimate of  the

linear term, q'J, consistent with the data, is calculated.  At low doses,  this

upper-limit linear term predominates, forcing the model  to low-dose linearity.



B.  USE OF THE LINEARIZED MULTISTAGE MODEL FOR RISK EXTRAPOLATION OF

    2,3,7,8-TCDD:  COMPARISON OF FOUR U.S. AGENCIES

     Besides the choice of the linearized multistage model for animal-to-human

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risk extrapolation, the final risk estimates are dependent on a choice of
several other factors.  Specifically, Table 2 presents a summary of  2,3,7,8-TCDD
cancer risk extrapolation by four U.S. agencies, all  of which used the linearized
multistage model.  Even with the use of the same model, however, the results
varied over 10-fold due to a selection of different factors relating both to
the animal data and to the procedure.  Such factors are:
          •  Choice of animal bioassay
             Adjustment made for differential nontumor mortality
             among the animal treatment groups
          •  Selection of tumor types for modeling
          •  Animal-to-hum an dose equivalence
          •  Dose used for curve fit
     As seen in Table 2, EPA, FDA, and CDC all used the cancer response data
from the female Sprague-Dawley rat in the 2-year feeding study conducted  by
Kociba et al. (1977,  1978).  In the EPA HAD for PolychloMnated'Dibenzo-j>-
Dioxins, the choice of the Kociba study was based on  the female rat  providing
the largest slope factor, q^, of all the available data sets.  However, there
were other, unstated but probably better, reasons for the selection, such as
(1) the high quality of the study, (2) response at multiple sites, (3)  more
applicable route of exposure to the human experience than the gavage study, and
(4) less controversial  tumor sites than the mouse liver.  The State  of California
used the liver tumor response from the male mouse in  the National  Toxicology
Program (1982)  gavage study and estimated an upper-limit incremental  unit risk
estimate of qj = 1.5 x 10~7 (fg/kg-day)"1, nearly the same as that of EPA
(qj = 1.56 x 10"7).  Also, in its analysis EPA contracted with an  independent
pathologist, Dr. Robert Squire, to provide a second examination of the liver
slides in the Kociba  study.  Even though Squire's analysis indicated more liver

                                       10

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tumors  in the control and low- and mid-dose groups, the estimates of an incre-
mental  increase in cancer risk differed by less than 10%.
     Another factor of concern in the extrapolation procedure was nontumor
toxicity among the female rats.  In order to correct for high early mortality
in the  high-dose rats, EPA's analysis eliminated all animals that died during
the first year of the study (before the appearance of the first tumor).  The
elimination of nine animals in the high-dose group, one in the controls,  and
one in  each of the other dose groups, changed the upper-limit estimates by a
factor  of either +1.7 or 1/2.5 depending on which pathologist's analysis  was
used.   Since EPA was the only agency to make the adjustment, its estimate
incorporating both pathologists1  analyses actually decreased by a factor  of 2.7
compared with the unadjusted analysis.
     In selection of tumor types, all the agencies modeled the liver tumor
response.  EPA also included the cancer response in the lung and hard palate/
nasal turbinates, but this led to only a minor increase in the  final  estimate
since the liver produced the major response.
     Animal-to-man dose equivalence factors are discussed in the HAD.   Both  EPA
and the State of California used dose/surface area equivalences between animal
and humans.  The FDA used dose/body weight which reduces human  risk estimates
compared to surface area by a factor of 5.4 for rat-to-human extrapolation.
The CDC used liver concentration at terminal sacrifice, a measure that would
be preferable if human tissue distribution was also known.   In  the  present
case, however, the known rat liver concentration measures of dose equivalence
had to be equated back to the rat administered dose without  a comparable  known
relationship in humans.  The dose used for the curve fit by EPA,  FDA,  and the
State of California was the dose  actually administered  to the animals.  The
CDC's use of liver concentrations at terminal  sacrifice resulted in the risk
                                       12

-------
estimate being increased by a factor of 2.   Species  conversion  factors are
discussed further in section II.C.
     The end result of all  of these factors in the risk extrapolation procedure
resulted in a maximum difference of a factor of 9, that being between the EPA
and the FDA.  This can be seen either by comparing the upper-limit  incremental
unit risk estimates or the  risk specific doses (RsDs), which are just the
reciprocals xlO~6.

C.  ALLOMETRIC AND BODY BURDEN CONSIDERATIONS
     The dose metric or allometric  equivalence for rat-to-human extrapolation
has a potentially large quantitative impact on 2,3,7,8-TCDD risk estimation
because of the large differences in half-lives in the rat  and human.  However,
what little attention this topic has received from regulatory agencies until
now has taken into account  only the standard dose metrics.  As  shown  in  Table  2,
both EPA and the State of California used the administered dose/surface  area
conversion, the FDA applied an administered dose/body weight and the  CDC applied
the actual rat liver concentration  at terminal sacrifice.   When extrapolating
from rat to human, use of the dose/surface area allometry increases the  risk
estimate by a factor of 5.4 versus either dose/body weight or  liver concen-
tration.  When extrapolating from the smaller mouse to human, the corresponding
use of dose/surface area allometry results in a factor of 13 greater  risk
estimate.
     The EPA has used the dose/surface area metric, often called  a  species
extrapolation correction factor, as a conservative, prudent policy.  It  is  based
on the observations that among different mammalian species'many physiological
rates, and especially ventilation, basal metabolic, and clearance  rates  tend  to
scale  in proportion to a fractional power of body weight.  It  has  also  been found

                                       13

-------
  to hold for the acute therapeutic  effects of anticancer agents.  That fractional
  power, often between 0.6 and 0.8 is very close to the fractional power of 2/3
  relating  the surface area of cylindrical or round objects to their-volume.
  Since the density of most mammalian bodies is about the same, mass or body
  weight can  be  used  instead of volume, and hence the term surface area or (body
  weight)2/3  correction.   In simplified terms the allometry of basal  metabolism
  is  often  explained  by the observation that the amount of calories a warm blooded
  animal will  consume  is enough to maintain body temperature and  that loss of
  heat  is related to  surface area and not mass.   However,  the allometry of species
 conversion for carcinogen risk  assessment  is  far  more complicated than simple
 basal metabolism.  Even assuming that  the  basic mechanism  of the carcinogen
 stays the same from high- to  low doses,  there  are often  large species differences
 both in tissue distribution  and in  metabolic pathways to  form the active carcino-
 gen.  Often, it is not known whether the parent compound or  one  (or more) of
 its metabolites is the active carcinogen.  Almost never  is there a good  under-
 standing of  the mechanism.
      It is just because of these many  unknowns that regulatory agencies have
 been forced  to  adopt  a general default position of a surface area or body weight
 or parts per million (ppm) in air species conversion  factor.  The EPA most
 often  uses surface area, but sometimes uses ppm in air as a species dose
 equivalence,  based on  the known cross-species allometry for Q2 consumption.
 The  FDA position is to use dose/surface area allometry when the active carcinogen
 is a metabolite of the administered compound and to  use dose/body weight  when
 the active carcinogen is thought to  be  the  administered compound  itself.   Their
 reasoning for the latter case,  of which they consider  2,3,7,8-TCDD an example,
is that if the administered compound does not have to  be metabolized  to be
carcinogenic, then strict  dose/body  weight  considerations  should  apply.   The
                                       14

-------
CDC apparently agrees.  The EPA counters,  however, that-even  if the parent
compound is the active carcinogen,  its  activity  is related to its time in the
body, which in turn is related to  clearance time and  hence to dose/surface area
allometry.
     At this point, it may be instructive  to derive the dose/surface area allo-
metry in a slightly more rigorous manner.   The rat-to-human species correction
factor of 5.4 means that if the human were to receive the dose/body weight of a
compound, 5.4 times that of the rat,  the different elimination capacities of the
two species would cause both species'  concentration x time exposures to the
compound to be equal.   In terms of  first-order elimination kinetics for a
single dose of a nonmetabolized compound,  a rat  given concentration C0 with an
elimination constant ke would have  a  total  area  under the concentration-time
curve of C0/ke.  A human given a concentration of C0/5.4 would eliminate the
material, if allometric considerations  hold, at  a rate of ke/5.4, so that his
total  area under the concentration-time curve would be equal to that of the rat.
     For continuous daily exposure, the total area under the  concentration-time
curve is (C/k^) (TK - 1 + e~kT)5 where C  is the  daily dose/body weight and T and k
are units of days and reciprocal days,  respectively.   If T is large, say 730 days
                                    V
for a 2-year rat study and k is not very  small,  then  this area becomes approxi-
mately CT/k.  Thus, in order for the total  areas'to be the same for a 2-year
rat and 70-year human dosing period,  the  human would  have to be given a concen-
tration C/(5.4x35) or 1/189 that of the rat.  EPA's position  in this metric is
that one rat year is equivalent to 35 human years in  the cancer development
process, and that the cancer age-distributions for rats and humans are alike
when T is viewed as representing a lifetime.  Therefore, over a lifetime a
human should be allowed 35 .times the C/k  that of the  rat as an equivalent dose.
     Since rats and humans seem to follow closely enough for most compounds,

                                      15

-------
the clearance rate dose allometry discussed above, adjusting the species
correction factor for actual clearance rates, is not often  done  in  risk extrapola-
tion.  Furthermore, consideration of all the unknowns of actual  mechanism  and
metabolism make it apparent that the species correction  factor is only a rough
approximation, meant to somehow convey the concept of increased  sensitivity  of
the human compared to the smaller animals.  Still, the appropriateness of  the
use of the surface area correction factor is not clear in the case  of 2,3,7,8-TCDD
because of its extremely long half-life in humans compared  to rats.  The potential
effect of this difference on quantitative risk estimation is now examined.
     Rose et al. (1976), examined the fate of 2,3,7,8-TCDD  following single  and
repeated oral doses to Sprague-Dawl ey rats.  For a single oral dose of 1.0 ug
TCDD/kg body weight, they found a half-life, assuming a  one compartment open
model, of 31 +_ 6 days.  For repeated oral doses of 0.01,  0.1, or 1.0 ug TCDD/kg-
day, 5 days a week for 7 weeks, they found a half-life of 23.7 days.  For  the
single dose, after 22 days nearly all of the compound had concentrated in
either the fat or the liver, with equal  concentrations in each.   For the rats
administered the repeated oral  doses, the compound again  concentrated mostly in
the liver and fat, with the liver concentration being three to five times  as
high as that of the fat at the  end of 7 weeks.  This observation is consistent
with that of Kociba et al. (1978) who, in his 2-year feeding  study, found  liver
concentrations three to five times as high as adipose tissue when the daily  dose
was at least 0.01 ug/kg-day and about the same as the adipose tissue concentration
when the daily intake was 0.001 ug/kg-day.  Rose et al. estimated the elimination
constants for liver, fat, and whole  body all  about equal, 0.026  days"1, 0.029
days"-'-, and 0.029 days"1, respectively,,  corresponding to  half-lives of 24 to 27
days.  The relationship between half-life and clearance  times for first-order
kinetics is ti/2 = In2/ke.
                                      16

-------
     In humans the half-life of 2,3,7,8-TCDD  in  the  body has been variously
estimated as 3 to 5 years,  6 years,  10 years, and  if  2,3,7,8-TCDD acts according
to two compartment kinetics with the fat acting  as a  "deep" second compartment,
up to 30 years (U.S. EPA,  1988).  The CDC (1987) estimates a half-life of 6 to 10
years; this estimate will  be used here.
     An additional complication is that nonhuman primates, unlike rats, appa-
rently accumulate a higher concentration of 2,3,7,8-TCDD in the adipose tissue
than in the liver, with ratios ranging from 10:1 to  67:1.  The very  limited
human data also suggest an adipose tissue to  liver concentration ratio of 10:1,
with minor deposition in other organs.  One experiment exposing  rats and both
infant and adult monkeys to a single intraperitoneal  injection (400  ug TCDD/kg
body weight) found that after 7 days the rat  had concentrated 43% of the admin-
istered dose in its liver versus only 10% and 4.5% for adult and  infant monkeys.
In monkeys, the larger percentages were found in adipose tissue  (U.S. EPA,
1985).  Thus, if  the liver  is the organ of primary concern, for  tissue distri-
butions alone a human would have to be given  anywhere from 10 to 50  times the
dose on a mg/kg-body weight basis to have the same liver  concentrations as  the
rat.   If one is not concerned with the liver alone but with total body burden,
it  is  not these figures but the relative body half-lives  which  would apply.
     Estimates of the ratio of half-lives in the human versus the rat show  that
for a  human half-life of 2,190 to 3,650 days (6 to 10 years)  and  a  rat  half-life
of  approximately  25 days, the ratios are 88:1 to 146:1, far  higher than the
5.4:1  correction  used by  EPA.   If liver  is the focus and  comparative liver
tissue distributions are factored in, however, the rat-to-human  correction
factor ranges  from  1.8(88/50) to 36.5(146/4).
      The quantitative risk  implications of these adjustments  for  tissue distri-
bution and  half-life differences between the rat and the  human  are  presented

                                        17

-------
 in Table 3.  As can be seen, if total body burden (area under the  time-concentra-
 tion curve) only is considered, the very long half-life in  the human  leads to
 risk estimates between 16 and 27 times that in EPA's HAD (1985).   If  liver
 concentrations are considered however, the relative risks  range from  0.3 to 6.8
 times that of EPA's estimate.  All estimates are higher than  the  FDA's.  The
 limited evidence suggests that if liver tissue concentration-time  species
 equivalence is correct, then the EPA HAD (1985) might underpredict the upper-limit
 risk by a factor of 1.6 to 6.8.
D.  OTHER EXTRAPOLATION MODELS
1.  Longstreth and Hushon (1984)
     Alternative models have been used for extrapolating  to  low-dose  risk.
Three of these, the one-hit, the Probit,  and the  Weibull,  have been discussed
and modeled in Appendix C of the HAD (U.S. EPA,  1985).  The  latter two, plus
two others, the Logit and the multihits have been modeled  by  Longstreth and
Hushon (1984), but they have not adjusted their  data for  high early nontumor-
related mortality.  This has been done in Tables  4 and  5  for  the  Kociba and
Squire histopathology analyses, respectively. The resulting  MLE  and  upper-limit
risk estimates for several  low-dose levels are consistent  for both data sets.
     For both data sets the linearized multistage and one-hit models  yielded
identical results.  Also, for both data sets the  upper-limit  estimates based on
the multistage model  were consistent, while those based on the other  three
models varied considerably.  For example, at a dose level  of  10-5 ng/kg-day the
UCLs for the multistage model  were 1.5 x  10-6 and 1.6 x 10-6  for  the  Kociba and
Squire pathology analyses,  respectively.   However, at the  same dose level of
10~5 ng/kg-day, the UCLs for the Logit model varied by  a  factor of 29, from
2.0 x 10-6 for the Kociba pathology to 5.8 x 10-5 for the  Squire  pathology.

                                       18

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Also, at the same dose level, the UCL from the Weibull model  varied by a factor
of 13, while those based on the log Probit model  varied by a  factor of over 50
million.  In general, the UCL risk estimates based  on  these models exhibit the
relationship:

        Weibull > Logit > linearized multistage = one-hit »  log Probit

     All of these models fit the data satisfactorily in the animal experimental
range, yet as can be seen from Tables 4 and 5, the  estimates  can vary over
several  orders of magnitude at lower environmental  doses.  The choice of the
model must rely on factors other than goodness of fit.
2.  Sielken (1987)
     A second risk modeling analysis of the female  rat liver  response in the
Kociba et al. 2-year feeding study has been published by Sielken (1987), whose
arguments have also been reproduced by Paustenbauch et al. (1986).  Sielken
fits the multistage model to the data but  focuses on the large difference in
low-dose behavior depending on whether or  not the high experimental animal dose
of 0.1 ug/kg-day is included.  Si el ken's measure  of risk  is the VSD which he
defines in terms of 10~6 lifetime incremental  risk.  The VSD  is derived from an
extrapolation model of the MLE of the multistage  model and not from the upper-
limit estimate.  Sielken, thus, uses Crump's reparameterization of the multistage
model but not Crump's linearized form.
     Si el ken's analysis is based on the argument  that  use of  the multistage
model with the 0.1 ug/kg-day dose-response included forces the estimate of the
linear term to be positive non-zero and distorts  the true shape of the dose-
response curve.
     In particular, he argues, even though the multistage model fits the data
with the high-dose point included, "the (resulting) fitted models do NOT [his

                                       22
-------
emphasis] reflect the observed behavior at  the  lower  experimental doses."  A
proper low-dose shape, he claims,  is  seen when  the highest dose is removed.  In
this case, the linear term of the  multistage becomes  zero and the quadratic
term becomes positive.  Extrapolation with  the  quadratic curve goes rapidly to
zero compared to low-dose extrapolation with a  linear model.
     An examination of the results derived  from the type of analysis suggested
by Sielken can be seen in Table 6.  In this table, both the MLE and 95%  lower-
bound estimates of the VSD are calculated using different permutations of the
Kociba female rat data (animals dying before the  appearance of the first tumor
have been eliminated).  In every case, the  model  with estimated parameters fit
the data satisfactorily.  The first row contains  the  observed  liver tumors for
all doses and parameter estimates, while the second and Tower  rows omit  the
high-dose group.  The third through the seventh rows  permutate the low-dose data
by increments of one tumor-bearing animal.   The proportion  of  6/48 represents
the 95% upper-limit on the observed proportion  of 3 tumor-bearing animals out
of 48.
     An  examination of Table 6 shows the instability  of the MLEs  under this
model.   As was pointed out above, omitting  the  high-dose  group changes the form
of the model  from  linear  to quadratic, with a corresponding 330-fold  increase
in the VSD  from  4.8 x 10~8 to 1.6 x 10'5.  The  form of the  model  remains quad-
ratic until the  number of tumor-bearing animals is incremented by  2.   However,
when  the increment  becomes 3, to  total 6 out of 48,  which is the 95%  upper limit
of the actual  observed response,  the  picture again changes.   When this 95%
upper-limit of the  observed  low-dose  response  is fit, the model  again incorpor-
ates  a linear MLE,  and the MLE of the  VSD  returns to 5.0  x 10"8.
      In  contrast to the  instability of the MLEs, the 95% lower limits of the
 VSDs  remain quite stable over the range of permutations, varying by a factor  of
                                       23
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2.  The ratios of the VSDs  to their lower  limits  reflect the instability of the
MLEs for low-dose extrapolation  with  the multistage model.
     Si el ken's results are consistent with the  findings of  Portier and Hoel
(1983), who concluded that  for a multistage model  fit of the data, the MLE of
the linear term is usually multimodel, the number of modes  being  equal to the
degree of the chosen polynomial.  This instability of the MLEs from the multistage
family of models has also been confirmed  by several other authors; it  is further
seen in Tab! e 6.
     Crump (1987) provided a critical review of Sielken's analysis and observed
that even if "the true shape of the dose  response is a  straight line connecting
the background  response and the response  at the mid-dose,"  the probability of an
MLE estimate of zero for the linear term  in the multistage  model  is about  1/3.
He concluded that while the data are consistent with Sielken's interpretation
of a higher RsD, they are also consistent with those much  lower,  as displayed in
the confidence limits.
     It  was based on these types of analyses and a consideration  of the  typical
animal bioassay data to be fit that Crump had originally advised  that  an upper-
limit  estimate  of the low-dose  risk be used for his model.   Under his  reparameter-
ization  of the multistage model, an MLE of zero for the linear term becomes
possible,  and this  can cause great instability in low-dose risk  estimation.
However, under  the  original development of the model,  it  is not  possible to  have
higher degree polynomials without  having  a positive linear or first-stage  term.
Crump's  reparameterization is necessary with quantal data  in order to  limit  the
number of  parameters.   In doing  this, however, the MLEs can become unstable, as
is  seen  above.
                                       25
-------
E.  TIME-TO-TUMOR ANALYSES
     An extension of the multistage model analysis can be conducted when time
to observation of the tumors is known.  This extension is modeled  into  the
formulation of the multistage model:
P(d,t)  = l-exp-(q0
q2d
                                                     q3d3tk)
where qg, qi , q£ , Q3 » and k are parameters estimated from the data and are all
constrained to be nonnegative.  This model is often called the Weibull  model,
but is more appropriately described as the multistage Weibull , since  it is
multistage in dose but Weibull in time.  Its generalization over the  multistage
model  allows an estimation of the probability of cancer by a fixed age in  the
absence of any competing risks.  Its superiority over the quanta!  or  multistage
form is in its ability to adjust for treatment group differences in nontumor-
related mortality, as is seen in the Kociba study.  However, the analysis
requires a pathology decision as to whether the tumors of interest were fatal
or incidental, a condition not available in the Kociba study pathology.
     The results of a multistage Weibull model  are presented in Table 7 and  are
compared with two quanta! analyses using the linearized multistage model .   In
the first quantal approach no adjustment is made for the high early mortality
in the high-dose group.  In the second approach all animals dying  before the
appearance of the first liver tumor were dropped from the analysis (see section
II. B).  These two analyses are compared with the multistage Weibull under  the
extremes of either all  fatal  or all incidental  tumors.
     As can be seen from Table 7, the largest difference in risk estimates is
between the unadjusted analysis and the fatal  tumor for this analysis;  8.4
(for the MLE term) and 10 (for the upper limit).  However, both these extremes
are thought to be somewhat misrepresentative of the data, and either  of the

                                       26
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middle two approaches is considered superior.  While the present EPA analysis
(see section II.B) yields the lower of these estimates by a factor  of 2,  this
difference is considered minor.

                         III.  TREATMENT AS A PROMOTER

     Evidence for the promoting action of 2,3,7,8-TCDD in the rat liver  has
been well documented in the HAD (U.S. EPA, 1985).  This section compares the
quantitative cancer risk estimates under the assumption that 2,3,7,8-TCDD's
action on the liver is one of promotion.  It is shown that even when treated
solely as a promoter the estimates of risk can vary greatly according to  assump-
tions that one  is prepared to make.  In section III.A. the treatment of  2,3,7,8-
TCDD as a classical toxicant or promoter with a threshold is presented,  along
with the results of Canadian and several European regulatory agencies who
estimate "virtually safe" levels by applying uncertainty factors to dose levels
at which no adverse effects are observed.  In section III.B. the results of  a
new approach to modeling for promoters are discussed in which cancer response
is modeled as a function of liver cell proliferation of initiated cells.
A.  UNCERTAINTY FACTOR APPROACH
     Several countries and the State of New York have estimated RsDs  for
2,3,7,8-TCDD's potency by the application of uncertainty factors to NOELs,
NOAELs , or LOAELs .  The general approach is to use uncertainty factors ranging
from 10 to 1,000 based on a rul e-of-thumb approach (Cook and Page,  1986).
     •  A factor of 10 where adequate chronic human toxicity data as well  as
        adequate chronic oral toxicity data in more than one animal species are
        available;
                                       28
-------
        A factor of lO'O where adequate chronic  animal  toxicity data are avail-
                                         ,-
        able in more than one species, but  human  toxicity data,are lacking; and

     •   A factor of 1,000 where limited chronic animal toxicity data are avail-

        able (in only one species)  or inconclusive results  in more than one

        species.

     The National Academy of Sciences Safe  Drinking Water Committee was more

explicit with regard to carcinogenicity studies  (NAS,  .1977).

     "1.  Valid experimental results from studies on prolonged ingestion by man

          with no indication of carcinogenicity.

                               Uncertainty  Factor = 0

     2.  Experimental  results of studies of human ingestion  not available  or

         scanty (e.g., acute exposure only).  Valid results  of long-term feeding

         studies on experimental  animals or in  the absence of human studies,

         valid animal  studies on one or more species.   No  indication of carcino-

         genicity.
                               Uncertainty  Factor'= 100

     3.  No long-term or acute human data.   Scanty results  on experimental

         animals.  No indication of carcinogenicity.

                              Uncertainty Factor  = 1,000"

     Others have sought to break down uncertainty factors  into components. Weil

(1972)  interpreted  the application of the uncertainty factor of 100 as a product

of a factor of 10 to account for differential  human sensitivities and  the

second  factor of 10 to extrapolate results  from animals  to  humans.  When cancer

became the end point of concern, others included  a third factor of 10 to raise

the total to 1,000.  The uncertainty factors,  the toxic  end  points, and the

RsDs used by several" agencies for the evaluation  of 2,3,7,8-TCDD  are  presented

in Table 8.' The table includes results from:the  agencies that have used the
                                       29
-------








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linearized multistage model  as  a  comparison  to  those  agencies using the uncertainty
factor approach.   As  can  be  seen, the  RsDs generally  segregate into two groups
with those in the lower potency groups all  using  the  NOEL approach.  The State
of New York, as well  as the  countries  of  Canada (plus the Province of Ontario),
Switzerland, and Germany, consider  the cancer study of  Kociba to establish a
NOEL at 1 ng/kg-day.   The Netherlands  also  uses the Kociba  study as its per-
tinent cancer study,  but  considers  instead  the  1  g/kg-day dose as a non-noxious
dose due to the slight increase in  liver  cell changes in the female rats at
this level.  All the agencies,  except  Switzerland, chose the uncertainty factor
approach to establish an  RsD.  Switzerland  actually estimated inhalation and
oral exposure and divided those exposures into  the  1  ng/kg-day NOEL to determine
a "safety factor" attributable  to those exposures.
     As can be seen in Table 8, potency ratios  seem  to  gather  in  factors of  10.
The RsDs for the five agencies  above the dotted line  span one order of magnitude
as do those for the agencies below the dotted line.   Separating  the higher-  and
lower-potency groups is a factor of 15.6 (10 * 0.64).  The  total  potency range
is  1,564.   All but one agency uses the Kociba data.

B.  MODELING AS A PROMOTER UNDER THE MOOLGAVKAR,  VENZON, AND  KNUDSON
    TWO-STAGE  MODEL   •.
      A  variation  of the multistage model has been developed by Moolgavkar,
Venzon  and  Knudson (M-K-V, 1979, 1981) which models cancer  as  a. two-stage  process
with  a  promotion  phase.   This model has been shown to predict very well  tumor
promotion  in the  mouse skin  (Chu et al., 1987).  A variation  of  this  model  has
been  developed by Dr.  Todd Thorslund  to extrapolate  from animals to provide
human risk  estimates  of  liver cancer  deaths  (see the Appendix to this paper).
 The model  considers  that the carcinogenic action of  2,3,7,8-TCDD is through its
                                       31
-------
 dose-response  relationship on the proliferation of initiated liver cells.  By
 including what is  known about the receptor-mediated mechanism involved, cell
 proliferation  is  itself considered a function of 2,3,7,8-TCDD's binding capacity,
 which can be shown to follow linear but saturable kinetics.  When these parameters
 are  factored into  the model, the cumulative probability function, P(x)  for dose
 x at a fixed time t, becomes

                    P(x) = 1-exp-M  CexpG(x)t-l-G(x)t]/62(x)
                    6(x) = 6(0) + [G(co)-G(0)][l-exp-Vx]

 where P(x) = the probability of a tumor with dose*,
         M = the background mutation rate proportional  to background tumor rate,
      G(x) = the liver cell proliferation rate of initiated cells associated
             with dose x,
      6(0) = the background liver cell  proliferation rate,
      G(«>) = the maximum cell  proliferation rate possible, and
         V = the parameter associated with the liver saturable kinetics of
             2,3,7,8-TCDD-receptor binding.
     Even though this model is termed a two-stage model, conceptually it is rad-
 ically different from the multistage model in several  respects.   The multistage
model, as computed by EPA, is a basic curve-fitting model with all  parameters
estimated from  the data.  The two-stage model  with promoti'on as  constructed can
actually be fit without the estimation of any parameter from the cancer bioassay
dose-response data.  For example, M and 6(0) can be estimated entirely  from the
 control data, V can be estimated from a separate experiment measuring 2,3,7,8-
TCDD uptake by  the receptor, and 6(°°)  can be estimated  by the incorporation
of thymidine into the nuclei following administration of a saturation level of
 2,3,7,8-TCDD.  In theory, then, with the exception of the use of control  animals
for the estimation of background rates, all parameters would be estimated

                                       32
-------
separately and a goodness-of-fit test would be a good measure of how well  the
model fits the actual  data.
     An extension of the above distinction between the multistage model  and
M-V-K two-stage model  is that the parameters derived from a bioassay based on
one  rat strain can be used to predict the response from a second strain  with
only the background rate M having to be estimated from- the control  group data
of the second strain.  This procedure was extended fr,om animal-to-hum an  extrapo-
lation where human liver cancer death rates in the United States from 1980 were
used to estimate the background human rates for M and G(0), and the values of
the other parameters estimated from the rat data were used to provide human
risk estimates.
     In addition to modeling liver cell proliferation as a nearly linear func-
tion of 2,3,7,8-TCDD receptor binding (called here, the negative exponential
form of the M-V-K model), a second form of the model 'results when the 2,3,7,8-
TCDD-induced cell proliferation rate'is assumed'proportional  to the product  of
cellular 2,3,7,8-TCDD levels and the number of 2,3,7,8-TCDD receptors.  This
results in a model for the induced cell proliferation rate which is log-logistic
in form.
     Both models are used to extrapolate from the animal to human cancer response.
The  results are presented in Table 9 which is reproduced from the Appendix.
While  both forms of the M-V-K model fit the observed data quite well, and  both
can  be justified on theoretical and some experimental ground, their use  for  low-
dose extrapolation leads to a wide variation in risk estimates.  For a lifetime
daily  dose of 0.1 ng/kg-day, or one order of magnitude below the animal  experi-
mental level, the estimates vary by a factor of more than 10,000.  Furthermore,
even though the  negative exponential form of the model results in a .linear low
dose-response relationship, the risk estimates are 2 orders of magnitude below

                                       33
-------
     TABLE 9.  ESTIMATES OF LOW-DOSE INCREMENTAL RISK TO HUMANS EXPOSED TO
               2,3,7,8-TCDD BASED ON FEMALE SPRAGUE-DAWLEY RATS3
  COMPARISON OF THE TWO FORMS OF THE TWO-STAGE MODEL WITH PROMOTION WITH EPA's
           RISK EXTRAPOLATION USING THE LINEARIZED MULTISTAGE MODEL3

Dose
(ng/kg-day)
ID'5
10-4
lO-3
10-2
10-1
1
Promotion'3
Negative
exponential
1.7 x 10-8
1.7 x 10-7
1.7 x ID'6
1.7 x 10-5
1.8 x 10-4
2.4 x lO-3

Log-
logistic
—
--
<10-13
8.8x10-10
8.8x10-7
9.3xlO-4
Linearized multistage model
EPA
(Upper confidence limit)
1.6 x 10-6
1.6 x 10-5
1.6 x 10-4
1.6 x 10-3
1.6 x 10-2
1.6 x 10-1






aTaken from Table 11 of the Appendix.

^Squire's pathology analysis adjusting for early mortal ity.
                                       34
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the upper confidence limit obtained by EPA,  which modeled 2,3,7,8-TCDD as a
complete carcinogen.  This difference between  the estimates with the linearized
multistage model and negative exponential  fo'rm of -the  promoter model is due
primarily to the low human background liver  cancer  rates compared to those of
the female rat.  However, these estimates  pertain only to human  liver cancer.
In its present form, the model  is target organ-specific from  animals to humans.
     Estimates with the negative exponential form of this promoter  model might
be considered as providing a conservative, on  the high side,  estimate of
induced liver cancer risk, since cell proliferation is modeled as a linear
function of dose.  However, not enough research has been done on the low-dose
properties of these models to characterize the variability  of the low-dose  risk
estimates.  While the upper-limit estimates  should  certainly  be  below those of
the linearized multistage model, further research needs to  be done  before these
models can be used for regulatory decision-making.

          IV.  COMPARISON OF ANIMAL PREDICTION WITH ACTUAL  HUMAN DATA

     With the exception of one study on 2,3,7,8-TCDD in Holmesburg, Pennsylvania,
in 1967, all  human  exposure data are derived from accidental  exposures  of
unknown quantity.   In the Holmesburg study,  2,3,7,8-TCDD  was  topically  applied
to volunteer  prisoners in 'total doses ranging from  0.4 ug  to  7,500  ug.   According
to testimony  and exhibits in EPA's 2,4,5-T cancellation hearings (Rowe,  1980),
doses  below  16 ug  did'not el'icft a chlo'rac'ne response, while  a  dose of  7,500  ug
did cause chloracne  in 8 out of 10 subjects.  This  high dose  of  0.05 mL of  a  1%
2,3,7,8-TCDD  solution  in 50/50 alcohol  chloroform solvent was applied  to one
square inch of the backs every other day for a month and  covered by a  nonocclusive
patch.  If  we can  assume  that the  subjects'  average age was 35,  a 25% absorption

                                        35
-------
 rate,  an  infinite  half-life, and lifetime (70 years)  average daily dose  (LADD)
 proportionality, one can estimate an internal dose of 2,3,7,8-TCDD that  caused
 chloracne to  be  in the 2 to 1,000 pg/kg body weight-day range.  If we  assume
 either a  5- or 10-year half-life with first-order elimination kinetics and with
 the  other assumptions the same, then the chloracne causing the LADD range  is
 unchanged  at  one significant digit.  These figures correspond to  an external oral
 dose of 4 to  2,000 pg/kg body weight-day.
     Human cancer risk estimates based on the Kociba  female rat feeding  study
 with 55%  absorption yield an wpper-limit risk estimate of 1.56 x  10~4  (pg/kg-
 day)~l.   Multiplying this upper-limit estimate by 4 to 2,000 pg/kg-day and
 adjusting  for the different absorption fractions yields an upper-limit lifetime
 incremental cancer risk of between 6 x 10~4 and 3 x 10~1 for hunans developing
 chloracne.  Only 10 of the Holmesburg prisoners were  exposed to the highest
 dose,  and  follow-up is unclear.  Nevertheless, even if their lifetime  projected
 increnental cancer risks  were as great as 0.3, and even if they were observed
 for  their  full remaining lifetimes, less than three additional  cancers would be
 expected.  Put another way,  with three additional  cancers expected, an observation
 of no  additional cancers would not be highly unusual  (p = 0.055).   Clearly,
 unless specific types of cancer were to appear in these tested  prisoners, no
 conclusions could be drawn.
     On the other hand, Tschirley (1986) in his review of the 2,3,7,8-TCDD
 literature displays 9 cohorts of some 599 subjects who developed  chloracne
 following 2,3,7,8-TCDD and phenoxy herbicide exposure.  This is reproduced as
 Table 10.  Of those cohorts, only the study of the 1949 accident  at the  Monsanto
 plant  in Nitro, West Virginia (Zack and Suskind, 1980), and the 1953 accident
at the BASF plant in Germany (Thiess et al.,  1982), have sufficient latent
 period and other information to allow a comparison to be made between observed
                                       36
-------
          TABLE  10.  EXPOSURE TO 2,3,7,8-TCDD FROM INDUSTRIAL ACCIDENTS
Date
Workers
exposed
Location
of accident
Remarks
 1949
250
 1953
 75
 1956
 1963
106
 1964
 61
1965-69
 78
Monsanto plant in
Nitro, WV
BASF plant in
Ludwigshafen
Rhone-Poulenc
plant in
Grenoble

NV Philips
plant in
Amsterdam
Dow Chemical
plant in
Midland, MI
Continuing leaks
in Spolana plant
near Prague
122 cases of chloracne  being
studied;  32 deaths vs.  46.4
expected; no excess deaths
from malignant neoplasms or
circulatory disease

55 cases  of chloracne,  42
severe; 17 deaths vs. 11 to 25
expected  (four gastrointes-
tinal cancers and two oat-cell
lung cancers); most common
injuries  were impaired  senses
and 1iver damage

17 cases  of chloracne,  also
elevated  lipid and cholesterol
level s in the blood

44 chloracne cases (42 severe)
of whom 21 also had internal
damage or central nervous
system disturbances; eight
deaths (six possible myocardial
infarctions); some symptoms
of fatigue

49 cases  of chloracne;  4 vs.
7.8  expected deaths; 3 cancer
deaths vs.  1.5 expected; one
a  soft tissue sarcoma

78 cases of chloracne; five
deaths; many of the 50 workers
studied for more than 10 years
have hypertension, elevated
blood levels of lipid and
cholesterol, prediabetes;
significant amounts of severe
liver and  neurologic damage
SOURCE:  Tschirley, 1986.
                                       37
-------
 and  predicted  risks.  These involved 1.77 chloracne cases'and e-ithe'r -a 26- or
 30-year latency period.  Quantitative cancer risk estimates based on these
 cohort experiences are calculated below.  However, the large uncertainty in the
 exposure estimates should be emphasized, and the assumptions used to derive
 these exposure estimates are necessary simplifications.
A.  THIESS ET AL.  (1982)
1.  Description
     At a factory  in Ludwigshafen, Federal Republic of Germany, in 1953,  during
the hydrolization of 1,2,4,5-tetrachlorobenzol  to 2,4,5-trichlorophenol ,  an
accident happened  exposing at least 70 persons.  These 70 as well  as 4 additional
persons who were only exposed "for a short time during" 1954 and 1955 were •  •
included in the cohort.  Of the 74 persons, 66 suffered chloracne or severe
dermatitis.  All 74 persons were successfully traced through 1979.  Of the-74
persons in the cohort, 21 had died during the 26 years of observation, just
slightly more than- expected in any of five different control groups.   However,
there were seven cancer deaths observed versus 4.03 to 4.35 expected in these
control groups.  Of these seven cancers, three were stomach (IOD 151), one was
colon (ICD 153), and three were lung (ICD 162).  All seven occurred at least  10
years after the accident.  A 10-year latent period will  be assumed.  The  results
are presented in Table 11.   The control group represents the expected deaths
based on the mortality rates of Rhinehessia-Palatinate 1970-75.   The mortality
for stomach cancer was statistically significant (p = 0.016), while that  for
lung cancer was marginally  significant (p = 0.09).
2.  Exposure Estimates*
     Although there were no concurrent estimates or measurements of
*Contributions to this section were made by Drs. Jerry Blancato  and  Lorenz
 Rhomberg of the Office of Health and Environmental  Assessment.
                                       38
-------
    TABLE 11.   OBSERVED AND EXPECTED DEATHS FROM STOMACH,  COLON,  AND  LUNG  CANCER
            IN THE  74 CASES WITH  CHLORACNE  OR  SEVERE  DERMATITIS FROM  THE
                          1953 BASF PLANT IN LUDWIGSHAFEN
                (26-YEAR FOLLOW-UP WITH  AT  LEAST 10 YEARS'  LATENCY)
Cause of death   ICD No.    Observed    Expected     SMR    p-value
                                                   95%  Upper
                                                   confidence
                                                   1imit  based
                                                   on Poisson
                                                   distribution
Stomach cancer    151
                      0.52
          5.76     0.016     1.15  -  16.8
Colon cancer
153
0.24
                                                  4.17     0.21
                   0.05 -  23.2
Lung cancer
162
1.05
2.86     0.09
                                                                     0.57  -   8.3
Person-years at risk = 972..6.
                                       39
-------
 2,3,7,8-TCDD exposure, Rappe (1987) reported a mean level  of 100 pg/g in  adipose

 tissue of four exposed BASF workers some 30 years after exposure.   If we  are

 willing to make several simplifying assumptions, then we can estimate a  range

 of exposure.  The assumptions are:

      •  The body is treated as a one-compartment open model.

      •  A range of half-lives is assumed to be between 5 to 30 years.

      •  Adipose tissue-in a 70-kg human is about 2'0% or 14 .k'g.  ! .

      •  Absorption into the body is assumed to be: between^ 5% and 25%.  This

        figure is arbitrary and is chosen because there is less  than  55%  absorp-

        tion in the rat feeding studies.

      Under these assumptions the pertinent equation is:
                                 C(t) = C0 e
                                            -kt
where C(t) =  the concentration in the fat at the time of analysis,

        C0 =  the concentration in the fat immediately following  absorption  in
              1953,

         k =  the total body elimination rate constant, assumed to  be  the  sum
              of the elimination rate constants by various physiological
              processes, and

         t = the time since absorption was completed, assumed  to  be  30 years.

Further, we note:

                                 k = 0.693/t0>5


where tg.5 = the half-life of elimination from the body, and


                                Exposure = C0V/r


where V = the weight of the fat compartment (14 kg)  and r = the absorption

fraction.  In this case, the weight of the fat compartment is  used  to  calculate

the dose.  The other organs may be neglected because of the propensity of
                                       40
-------
2,3,7,8-TCDD to partition into  the  fat.   For  example, the concentration in the
fat is about 100 times  that  of  the  blood.
     Based on the above assumptions and  estimates, the  range of estimates of
2,3,7,8-TCDD exposure  in  the BASF plant  accident is between 10 and 1,800 ug.
Based on a LADD, these estimates  range from 1.5 to 50 pg/kg body weight-day.
The calculations are shown in Table 12.   Clearly, this  factor of 180 in the
range of exposures and a factor of  30 in the  range of LADDs creates significant
uncertainty in the risk estimate  calculations.   In order to narrow the range,
however, we note that in the Holmesburg  study exposures below 16 ug caused no
chloracne, while exposures of 7,500 ug caused 80% chloracne.  Considering the
probably greater absorption in  the  Holmesburg study, the exposure estimates in
the top three rows of Table 12, showing  the  range of 220 to 1,800 ug, seem the
most likely.  We therefore adopt  LADDs in the range  of  6 to 5.0 pg/kg body weight-
day.  The risk estimates below will be calculated  on the LADD of 50 pg/kg/body
weight-day; risk estimates based  on the lower end of the range would be about
eight times higher.
3.  Risk Estimates
     Cancer risk estimates are calculated for the  stomach  cancer and lung cancer
mortality presented in Table' 11.   These estimates are presented in Table 13.
Two models  are  considered, the additive and  relative risk  models, which have
been used in several EPA risk assessments.  They  are developed and more fully
explained  in the recent  EPA update on dichloromethane  (U.S.  EPA,  1987).  Both
models  require  estimates of LADDs,  and these have  been  estimated above as
between 6  and 50 pg/kg body weight-day.  The results show that  incremental
cancer  risk estimates based on these human data are  higher than those based on
the  animal  data in  every case.  If the lower end of the range of  LADDs  had  been
used, the  human estimates would have been greater still.   The conclusion based

                                       41
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on the  above  analysis is that the upper-limit incremental  unit risk estimates
based on the  animal studies do not overestimate the human  risk from 2,3,7,8-TCDD
if either the lung or stomach cancer mortality response in humans is bona  fide.
B.  ZACK AND SUSKIND (1980)
1.  Description
     In Nitro, West Virginia, on March 8, 1949, excessive temperatures  in  an
autoclave involved in the production of 2,4,5-trichlorophenoxyacetic  acid
caused a relief valve to open, allowing fumes and residues to escape  into  the
atmosphere and into the interior of the building.  A total  of 121  white males
were identified as having developed chloracne following this incident,  and
these were included in the subsequent follow-up study,  with vital  status ascer-
tained through the last day of 1978, nearly 30 years.
     There was 100% follow-up of subjects;  89 were still  alive and "32  were
verified deceased by death certificate."  With 46.41 expected deaths, the
standardized mortality ratio (SMR)  for all  causes, 69,  was  significantly
(p < 0.05) lower than expected.  The only cause of death that displayed normal
rates was cancer which had 9 observed and 9.04 expected,  SMR=99.6.  Of  those
cancer deaths, five were from lung  (2.85 expected, SMR=175), three were from
lymphatic or henatopoietic tissue (0.88 expected, SMR=341), and  one was a  soft
tissue sarcoma (STS) (0.15 expected).  Only one of the  cancer deaths, a lung
cancer, was a nonsmoker.
2.  Risk Estimates
     In order to make any kind of quantitative risk estimation of  the potency
of 2,3,7,8-TCDD, several assumptions must be made.  These are:
     (a)   Exposure.  As discussed above for the Holmesburg  study,  it  is assumed
that the LADD necessary to cause chloracne was in the 2 to  1,000 pg/kg-day range,
                                       44
-------
An estimate in the middle, or 500  pg/kg-day,  appears to be a reasonable starting
point, as does a value of 150 pg/kg-day,  which  is  close to the geometric mean.
The value of 500 pg/kg-day will  be chosen as  the LADD dose for those who devel-
oped chloracne, but the uncertainty of this estimate must be stressed.  This
dose estimate is 10 times greater  than that estimated  for the BASF study; the
incremental unit risk cancer estimates will be  correspondingly lower.
     (b)  Expected deaths from cancer.  Table 14 presents the observed and
expected cancer deaths for the 29.8-year latency.   However,  all the cancer
deaths appeared after at least a 10-year latency,  and  it  seems  reasonable that
10 years is an appropriate latent  period for  any  2,3,7,8-TCDD-related cancer
to  express  itself.  Therefore, the expected  deaths presented by Zack  and Suskind
must be adjusted by subtracting the first 10  years'  experience.   An examination
of  vital statistics rates for lung cancer and STS suggests  that for lung cancer
deaths approximately 20% of a 30-year death  experience happens  in the  first  10
years; for STS the  figure  is approximately 30%.  Based on these adjustments,
the expected  deaths for this exposed cohort  become 2.3 and  0.10 for lung and
 STS cancers,  respectively.
      (c)   Person-years  at  risk.   A figure needed for the additive risk model
but not  the relative  risk model is the person-years at risk.  Since  the  first
 10 years are considered to  be a latent or risk-free period, only the last  19.8
years  are  counted  as  person-years  at  risk.   The follow-up was complete and
 there were 32 deaths.  It  can be  assumed that  the average time until  death was
 20 years from first exposure (for the  nine cancer deaths the average time  from
 the accident until  death was 22 years).   Therefore, the total person-years at
 risk can be estimated as
                        P-Y = 19.8 x 89 + 10  x  32  = 2,082
                                        45
-------
TABLE 14.  FOLLOW-UP OF 121 CHLORACNE CASES AT THE  MONSANTO  COMPANY  (NITRO,
      WEST VIRGINIA) USED TO DERIVE  QUANTITATIVE CANCER RISK ESTIMATES
Lung cancer
Unadjusted Adjusted
Deaths
Observed 55
Expected 2.85 2.3
Person-years
at risk — 2082
SMRs
Observed 175 217
95% Confidence 57-410 70-508
STS
Unadjusted Adjusted
1 1
0.15 0.10
2082
667 1000
9-3707 13-5560
   1imits  based  on
   Poisson
   distribution
                                    46
-------
     The results of the analysis using both the  additive  and multiplicate models
are presented in Table 15.   Based  on  the  assumptions discussed above, both
models yield similar results.  For extrapolation based  on lung cancer deaths,
the MLEs for incremental risks are 9.1 x  lO'2  and 4.5 x 1Q-2  (mg/kg-day)-1 for
the additive and multiplicative models, respectively.   For  extrapolation based
on the one STS death, the MLEs are lower  than  are those for lung,  3.0 x ICT2
and 8.6 x 10~3 for the additive and multiplicative models,  respectively.  The
95% upper-limit estimates presented in Table 15  are  those based  on asymptotic
normality theory and are, therefore,  lower than  would be  produced  by applying
the upper confidence limits on the SMRs from Table 14,  which  are based on the
Poisson distribution.  Those Poisson  upper-limit adjusted SMRs,  for  example,
would yield 95% upper-limit cancer unit risk estimates  of 5.2  x  10"2 (ng/kg-day)"1
for STS, and 1.6 x 10~i for lung cancer deaths for the  multiplicative model.
These values are significantly greater than those presented in Table 15  [1.4 x
10-2 and 1.1 x 10"1  (ng/kg-day)-1, respectively] and reflect  the high degree of
variability due to small sample size.
     Also presented  in Table 15 is.the quantitative cancer risk  estimate derived
from the Kociba feeding study (U.S. EPA,  1985).   While  the MLEs  based on the
animal data are si ightly higher than those based on the Monsanto data, the
differences are no greater than a factor of 2.4  for  lung  cancer  and 13.4 for
STS.  All the 95%  upper-limit estimates based on the human data  are within a
factor of 2 of the upper-limit estimate based on the rat  data.
     The conclusion  based on the above analysis  is similar to that derived from
the BASF analysis--the available human cancer data on 2,3,7,8-TCDD do not
provide  any evidence that the unit risk estimate based  on rat data overpredict
the human experience.  The information on human  exposure  is just too uncertain
to  allow for a more  definitive statement.
                                       47
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                           V.   DISCUSSION  AND SUMMARY

     This report has  presented  the  effects on the human cancer risk estimates
from 2,3,7,8-TCDD exposure under varying assumptions involved in the animal-to-
human extrapolation procedure.   It  has compared EPA's risk estimates with those
of other agencies, both U.S.  and foreign,  discussed the rationale, used in each,
and shown the effects of slightly different  assumptions on the estimates.  In
general, the risk extrapolations divide  roughly  into two  groups, those agencies
using the linearized  multistage family of  models for extrapolation and those
using an uncertainty factor approach. The agencies using the linearized multi-
stage model  all  produce RsDs  within a factor of  10; similarly, the agencies
using the uncertainty factor approach are  also with a  factor of  10.  The two
groups, however, are  separated  by a factor of 16, so that the lowest RsD, that
of EPA which used the linearized multistage  model,  is  1,600 times lower than
that of Health and Welfare Canada which  used an  uncertainty factor of 100.
     While EPA's cancer risk estimates were  the  highest of the agencies pre-
sented, other methodologies consistent with  the  data have yielded still higher
estimates.  For example, fitting the data  with  both the Logit and the Weibull
models would have produced significantly higher  estimates—the Logit by roughly
one  order of magnitude and the Weibull  by 2 orders  of  magnitude.  In addition,
even extending the multistage model to the Weibull-in-time model under a time-
to-tumor analysis would have increased the upper-limit estimates by  as much  as
a  factor of 5.  Of even greater uncertainty is  the  extremely  long half-life  of
2,3,7,8-TCDD in the  human compared to the rat.   If  half-life  is  related  to  species
sensitivity as  implied by the cross-species extrapolation factor, then recent
estimates of human half-life of 6 to 10 years implies  that rat-to-human  extrapo-
lation estimates  should be significantly higher, probably by  a  factor of  2  to 7.

                                       49
-------
      Finally,  a  new methodology has been presented that models 2,3,7,8-TCDD as
 a carcinogen promoter only.  The methodology models the promotion (liver cell
 proliferation) phase of the M-V-K two-stage model as a linear saturable function
 of  2,3,7,8-TCDD-receptor binding.  The risk estimates of induced human liver
 cancer  are  2 orders of magnitude less than those using EPA's current method-
 ology.   However, research on this new methodology is ongoing and, at this time,
 not enough  is  known about the low-dose estimation properties to make definitive  .
 statements.
      Which  of  these is the "correct answer"?  Probably "none of the above."
 What  the above analyses show are that all  the answers are consistent with the
 observed d-ata, and all have some credence depending upon the be! ievabil ity of
 the assumptions used.   The most pertinent fact is that 2,3,7,8-TCDD causes
 liver, tongue, hard palate/nasal turbinates, and lung tumors in rats at doses
 and conditions to which humans would never be exposed.  As  such,  even  with
 animal bioassays, as well-conducted as were those for 2,3,7,8-TCDD,  the informa-
 tion they contain for low-dose extrapolation i,s very limited.   Within  a  100-
 fold  decrease  from experimental  dose levels, the range of estimates predicted  by
models that fit the data well, rapidly diverge to a "pay  your  money,  take your
 choice" level   of 3 orders of magnitude.  Below that, divergence is  even  more
 rapid (see Tables 4 and 9).   Furthermore,  when extrapolation is made  from
 animals to humans, the uncertainty about the effect of the  extremely, 1 ong half-
 life in hunans gives concern  about  the conservativeness of the  upper-limit based
 on the surface area correction for extrapolation.
     Use of human data  for risk  assessment  purposes  is also  impossible with
2,3,7,8-TCDD.   First,  the evidence for human carcinogenicity of 2,3,7,8-TCDU
alone is judged inadequate (see  Appendix  B).   Second,  the studies providing  posi-
tive evidence  for carcinogenicity  are  of a  case-control design; these  lack both
                                       50-
-------
a population base and exposure estimates.   The  few cohort  studies available
lack sensitivity because of a  combination  of  factors  including exceptionally
small cohorts, insufficient latency,  and in some instances, little  evidence of
exposure.  Only with one study was  there even an estimate  of  2,3,7,8-TCDD
exposure, and that study lacked power to discredit any of  the predictions
provided by the animal  data.
     Well, the next question is whether any of  these  estimates can  be  con-
sidered superior to the others.  To answer this, one  must  first  presume that
mathematical models can be used for prediction, and then decide  on  whether to
model 2,3,7,8-TCDD as a complete carcinogen or  as a promoter  only.   Modeling as
a complete carcinogen, only the one-hit and multistage models have  a theoretical
backing in carcinogenesis; the other models presented are  merely well-known
tolerance distribution models.  The EPA position is that  use  of  the linearized
multistage model for extrapolating  upper-limits of incremental risk is both
prudent and protective.  When both animal  and human data  have been  available
for  risk estimation, the linearized multistage  model's use with  animal data has
provided estimates comparable with those derived from human data.   Furthermore,
low-dose supralinearity is rarely seen, so that using the  linearized multistage
model  for extrapolating from experimental levels probably represents a protective
level  for humans.
     When one models 2,3,7,8-TCDD as a promoter only, many additional  uncertain-
ties arise.  The two most  important for modeling are those associated  with
 reversibility  and  threshold.   Classical promoters are known to show reversi-
bility  of lesions  when the promoter is no longer administered and  cleared  from
 the system.   Furthermore,  large doses are typically required for promotion,
 indicative  of  a  toxicity effect either leading directly to cell  damage and
 regenerat-ion, or overwhelming  the cell's ability to prevent the  promoter  from
                                        51
-------
 reaching its site of action.   In  the  case  of  2,3,7,8-TCDD, the evidence for
 liver cancer in  animals  points  to  a mechanism of promotion via receptor binding,
 yielding a very  strongly bound  complex.  Even if this complex is broken down,
 the persistence  of  2,3,7,8-TCDD allows it  to bind with another receptor.  In
 terms of mathematical modeling, this  information translates to a dose-response
 function  for cell proliferation which excludes threshold and reversibility, but
 otherwise can be defined by Michaelis-Menten type kinetics.  When this function
 is  substituted into  the  promoter form of the M-V-K model, the results yield
 low-dose  estimates  of liver cancer for humans which are 2 orders of magnitude
 lower than the upper-limit estimates provided by the one-hit and linearized
 multistage models.
     While the upper-limit estimates can be considered upper-limits for total
 human 2,3,7,8-TCDD-induced cancer, the predictions with the promoter form  of
 the  M-V-K model  require  further examination.  First, they apply only to human
 liver cancer, a condition reported only in  the cohort exposed to dibenzofuran-
 contaminated  PCBs in Yusho, Japan (Amand et a!., 1984).  They do not include
 estimates for STS or non-Hodgkins lymphomas.  Second, they are considered
 "best" estimates compared with the upper-limit estimates calculated from the
 linearized multistage model.   Third,  the  form  of the M-V-K model  used  predicts
 carcinogenic  response only on the basis of promotion.  If a 2,3,7,8-TCDD-induced
 initiation stage  had been incorporated (as  is  suggested  by Holder and  Rosenthal ,
 1987) the low-dose cancer predictions would have been higher.
     For these reasons,  the estimates  provided by  the promoter form  of the
M-V-K model might be considered prudent lower  bounds on  cancer  risk, while those
upper-limit estimates provided by  EPA's current  methodology are to be considered
upper bounds.  While any number of assumptions could produce either  lower or
higher risk predictions,  the  biological facts  incorporated  into both models

                                       52
-------
provide what should be considered reasonable  working  limits.  In addition,
2,3,7,8-TCDD1s many other toxicities  should also be a consideration in setting
a lower limit.  Any criteria level  lower  than that provided by the M-V-K model
would be at a  level  where these  concerns  would prevail.  A final caveat remains,
however, on the impact of the extremely long  half-life of 2,3,7,8-TCDD in man.
If 2,3,7,8-TCDD is not sequestered  in the  fat, but is bioavailable, the
quantitative risk estimates would be  considerably larger.
                                       53
-------
                                 VI.   REFERENCES'
 Anand,  M.;  Yagi,  K.;  Nakajima  H.;  Takehora R.;  Sakai H.; Umed, 6.; Labor K.
      (1984)   Statistical  observations  about the causes,of the death of patients
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 Armitage, P.;  Doll,  R.   (1954)   The  age distribution of cancer and a'multistage
      theory of carcinogenesis.   Br.  J. Cancer 8:1-12.

 Barnes, D.  (1987)  Holmesburg prison-based assessment.  Unpublished notes.
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 California.   (1985, April  19)  Health  effects of 2,3,7,8-TCDD and related
      compounds (draft).   Department  of Health Services.  Berkeley, California.
      Unpublished.

 Centers for Disease Control  (CDC)  (1980)  Serum dioxin in Vietnam-era veterans.
      Preliminary report.   Morbidity  and Mortality Weekly Report 36(28):470-475.

 Chu,  K.C.; Brown, C.C; Tarone, R.E.; Tan, W.Y.  (1987)  Differentiating among
      proposed mechanism for tumor promotion in mouse skin  with the use of the
      multievent model for  cancer.  J.  Natl. Cancer Inst. 79(4)789-796.

 Crump,  K.S.   (1983)  How to handle animal  survival  data in  quantitative risk
      assessment: a discussion paper.   Prepared  for Research Triangle Institute,
      EPA Contract No. 68-01-6826, Delivery Order 28.

 Crimp,  K.S.   (1988)  A critical  evaluation of a dose-response assessment  for
      TCDD.  Food. Chem.  Toxicol.   In press.

 Crump,  K.S.; Watson, W.W.  (1979)  GLOBAL 79:  A fortran program to extrapolate
     dichotomous animal  carcinogenicity data  to low dose.   Prepared  for the
      National   Institute of Environmental  Health Sciences,  Contract No.  I-ES-2123.

 Crump,  K.S.; Guess,  H.A.; Deal,  L.L.   (1977)   Confidence intervals and  test of
      hypothesis concerning dose-response relations inferred from animal car-
     cinogenicity data.   Biometrics 33:437-451.

 Cook, B.T.;  Page, N.P.  (1986,  February 3)  Comparison of  carcinogenicity risk
     assessment of 2,3,7,8-TCDD (draft).   EPA Contract  No.  68-02-4131,  Work
      Order No. 3.

Crockett, P.W.; Crump, K.S.  (1986,  February)   Methods  for  assessment of  non-
     cancer  health risks.  Prepared  by  K.S.  Crump  and  Co. for  the  Electric
      Power Research Institute,  Contract No. RP 1826-17.

Environmental  Canada.  (1984)  Chlorophenol s  and their  impurities  in the  Canadian
     environment.  1983  Supplement.   Economic  and  technical  review report.
     EPS3-EP-84-3.   Environmental Protection  Programs,  Directorate,  Canada.
                                       54
                                                      i'
-------
Food and Drug Administration (FDA).  (1983,  June 30)   Statement  by S.A.  Miller,
     Director, Bureau of Foods,  FDA, before  the Subcommittee  on  Natural  Resources
     Agriculture Research and Environment, U.S. House of Representatives.

Germany.  (1984)  Report on dioxins.  Update to Nov.  1984'.   Federal  Environ-
     mental  Agency.

Howe, R.B.;  Crump, K.S.  (1982)   WEIBULL82:  a FORTRAN program for low-dose
     extrapolation using time to tumor data.  K.S.  Crump and  Company,  Inc.,
     Ruston, Louisiana.

Holder, J.M.; Rosenthal, S.  (1987, February 13)  Issue paper concerning the
     mechanism of 2,3,7,8-TCDD carcinogenicity.  Office of  Health and  Environ-
     mental  Assessment, U.S. Environmental Protection Agency, Washington,  DC.
     Unpublished draft report.

Kimbrough, R.D.; Falk, H.; Stehr, P.; Fries, 6.  (1984)  Health  implications
     of 2,3,7,8-tetrachlorodibenzo-_p_-dioxin  (2,3,7,8-TCDD)  contamination of
     residential soil.  J. Toxicol. Environ. Health 14:47-93.

Kociba, R.J.; Keyes, D.G.; Beyer, J.E.; et al .  (1977)  Results  of a two-year
     chronic toxicity and oncogenicity study of 2,3,7,8-tetrachlorodibenzo-_p_-
     dioxin  in rats.  Submitted to the U.S.  Environmental Protection Agency.
     Unpublished.

Kociba, R.J.; Keyes, D.G.; Beyer, J.E.; et al.  (1978)  Results  of a two-year
     chronic toxicity and oncogenicity study of 2,3,7,8-tetrachlorodibenzo-_p_-
     dioxin  in  rats.  Toxicol. Appl. Pharmacol. 46(2): 279-303.

Longstreth,  J.D.; Hushon, J.M.   (1983)  Risk assessment for 2,3,7,8-TCDD.
     In: Tucker, R.E.; Young, A.L.; Gray, A.P., eds.  Human and  environmental
     risk of chlorinated dioxins and related compounds.  New York, NY:
     Plenum, pp. 639-664.

Moolgavkar,  S.H.; Venzon, D.J.   (1979)  Two-event models for carcinogenesis.
     Incidence curves for childhood ad adult tumors.  Math  Biosci. 47:55-77.

Moolgavkar  S.H.; Knudson, A.G.   (1981)  Mutation and cancer: a model for
     human carcinogenesis.  J. Natl. Cancer Inst. 66:1073-1052.

Nati.onal Academy of  Sciences  (NAS).  (1977)  Drinking water and health.  Safe
     Drinking Water  Committee, National Academy of Sciences.  Washington, DC.

National Research  Council of  Canada (NRCC).  (1981)  Polychlorinated dibenzo-^-
     dioxins: criteria for their effects on man and his environment.  Pub!.  No.
     18474,  ISSN 0316-0114.   NRCC/CNRC Associate Committee on Scientific
     Criteria for  Environmental  Quality,  Ottawa, Canada.  251 p.

National Toxicology  Program  (NTP).  (1982)  Carcinogenesis bioassay of  2,3,7,8-
     tetrachlorodibenzo-jD-dioxin  in Osborne-Mendel rats  and B6C3F1 mice.
     Technical  Report  No.  209.   Research  Triangle Park,  NC.

Ontario, Canada,  Ministry of  the Environment.   (1985)   Scientific criteria
     document  for  standard  development.   No. 4-84.   PCDDs and PCDFs.

                                       55
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Paustenbauch, D.J.; Shu, H.P.; Murray, F.J.  (1986)  A critical examination of
     assumptions used in risk assessment of dioxin contaminated soil.  Regul.
     Toxicol. Pharmacol. 6:284-307.

Portier, C.; Heel, D.  (1983)  Low-dose-rate extrapolation using the multistage
     model.  Biometrics 39:897-906.

Rappe C.; Aldersson, R.; Bergquist, P.  (1987, January)  Sources and relative
     importance of PCDD and PCDF emissions.  Presented at the ISWA/WHO-EURO/DAKOFA
     Specialized Seminar on Emissions of Trace Organics from Municipal  Solid
     Waste Incinerators, Copenhagen, Denmark.

Rose, J.Q.; Ramsey, J.C. ; Wentzlin, T.H,,; Hummel, R.A.; Gehring, P.J.  (1976)
     The fate of 2,3,7,8-TCDD following single and repeated oral doses to the  rat.
     Toxicol. Appl. Pharmacol. 36:209-226.

Rowe, V.K.  (1980)  Direct testimony at U.S. EPA's cancellation hearings of
     2,4,5-T.  Exhibit 865.  FIFRA Docket Nos. 415 et al.

Sielken, R.L.  (1987)  Quantitative cancer risk assessments for TCDD.  Food
     Chem. Toxicol. 25(3):257-267.

Sielken, R.L.; Carlborg, F.W.; Paustenbach, D.J.; Shu, H.P.; Murray, F.J.
     (1986)  Alternative approaches to mathematically analyzing the bioassay
     data for 2,3,7,8-TCDD (Abstract 1133.)  Presented at the 25th Annual
     Meeting of the Society of Toxicology, New Orleans, LA.

Squire, R.A.  (1980)  Pathologic evaluations of selected tissues from the
     Dow Chemical  TCDD and 2,4,5-T rat studies.  Prepared for the Carcinogen
     Assessment Group, U.S. Environmental Protection Agency, Contract No.
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Switzerland.  (1982)  Environmental pollution due to dimins and furans from
     communal rubbish incineration plants.  Federal  Office of Environmental
     Protection, Bern.

Thiess, A.M.; Frentzel-Beyme, R.; Link, R.  (1982)  Mortality study of persons
     exposed to dioxin in a trichlorophenol-process accident that occurred  in
     the BASF on November 17, 1983.  Am. J. Ind. Med. 3:179-189.

Tschirley, F.  (1986)  Dioxin.  Scientific American 25(2):29-35.

U.S. Environmental Protection Agency (EPA).  (1980)  Risk assessment on (2,4,5-
     trichlorophenoxy)acetic acid [2,4,5-T], (2,4,5-trichlorophenoxy) propionic
     acid, and 2,3,7,8-tetrachlorodibenzo-jD-dioxin [TCDD],   Office of Health and
     Environmental Assessment, Washington, DC.  EPA 600/6-81-003.   NTIS
     PB81-234825.

U.S. Environmental Protection Agency.  (1985)  Health assessment document for
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U.S. Environmental Protection Agency.  (1986)  Guidelines for carcinogen  risk
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                                       56
-------
U.S. Environmental  Protection Agency.  (1987)   Chapters.   Epidemiology:  recent
     Kodak study.   In:  Technical  analysis of ne^ methods  and  data  regarding
     dichloromethane hazard assessments.  Office of Health and Environmental
     Assessment, Washington, DC.   External Review Draft.   EPA/600/8-87/029A.

U.S. Environmental  Protection Agency.  (1988)   Estimating  exposures  to 2,3,7,8-
     TCDD.  Exposure Assessment Group, Office of Health and Environmental
     Assessment, Washington, DC.   External Review Draft.   EPA/600/6-88/005A.

Weil E.S.  (1972)   Statistics vs. safety factors and scientific judgement  in
     evaluation of safety for man.  Toxicol. Appl. Pharmacol. 21:254-463.

Zack J.;  Suskind,  R.  (1980)  The mortality experience of  workers  exposed  to
     TCDD in a trichlorophenol process accident,  J. Occup. Med. 22(l):ll-rl4.
                                       57
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                       APPENDIX
           QUANTITATIVE DOSE-RESPONSE MODEL
       FOR THE TUMOR PROMOTING ACTIVITY OF TCDD
                     Prepared by:
               Todd W. Thorslund, Sc.D.
                ICF-Clement Associates
                         1850 K Street, N.W., Suite 450
                         Washington, D.C.  20006
For:  The Carcinogen Assessment Group
      (Dr. Steven Bayard Project Manager)
      Research and Development
      U.S. Environmental Protection Agency
      EPA Contract No. 68-01-6939
                   February 17, 1987
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 1.   TECHNICAL  SUMMARY








     A method is developed  in  this report  for  the estimation of



 cancer risk associated with exposures to TCDD  under the assump-



 tion that TCDD  acts exclusively as a tumor promoter.  It is



 recognized that TCDD may have  other potential  effects on the



 mechanisms of carcinogenesis;  however, evidence suggests that



 stimulation of  cell growth  by  direct or indirect mechanisms could



 play an important role in TCDD's carcinogenic  effect.








     TCDD's ability to affect  cell proliferation can be described



 using a mathematical dose-response model.  A very explicit defi-r



 nition of promotional activity can be used within the context of



 the Moolgavkar-Knudson-Venzon  two-stage model  to accomplish this.



 It is assumed that a promoter  can act by increasing the net



growth rate of  preneoplastic,  initiated stem cells and has the



 ability to increase the growth rate to a fixed upper bound.  The



difference between this upper  bound and the background growth



rate is the maximum increase possible.  The mathematical function



that defines the fractional part of the change in the maximum



 increases in the growth rate due to a given exposure level of an



agent is the critical element  in the derivation of a dose-response



relationship.    Information  that can'be used to elucidate the form



of the critical function that defines the exposure-dependent cell



growth rate comes from the  three sources:  (1)   theoretical biolog-



ical arguments,  (2)  the shape of tumor dose-response relation-



ships from TCDD carcinogenesis bioassays, and  (3)  studies of
-------
TCDD-induced cell proliferation rates as direct or indirect



measures of cell growth in vivo or in vitro.








     In this report, two separate parametric models describing



the dose-dependent changes in cell growth are postulated.  Each



model is consistent with a series of studies on the mechanisms of



action of TCDD.  The first growth rate model assumes that the



changes follow first order kinetics.  This assumption results in



a negative exponential model that contains only one parameter



requiring estimation.  This model is fitted to the most extensive



data set available  (liver tumors in female Sprague-Dawley rats)



in order to estimate the three unknown parameters in the tumor



dose-response model.  The validity of the resulting model is



tested in four different ways:  (1) its goodness-of-fit to the



tumor data, (2) its consistency with TCDD-induced proliferation



data, (3)  its ability to predict the TCDD-induced tumor response



in other sexes and;strains of the same species (rats) using



adjustments only for background tumor rates, and  (4) its ability



to predict the TCDD-induced tumor rates in another species



(mice), making adjustments for background rates and estimating a



separate growth rate parameter for that species.
     A second model is investigated that assumes the rate of



change of the TCDD-induced growth rate is proportional to the



product of cellular TCDD levels and the number of TCDD receptors.



The resulting model for the increased cell growth rate is log-



logistic in form.  This model has two parameters, the slope and
-------
intercept; only the latter can be estimated with the available



tumor dose-response .data.  In order to use this model,  the slope



parameter needs to be specified.  This is done by specifying



slopes that span the range of those determined from other types



of biological systems that are log-logistic in form.  The



intercept associated with a slope is then estimated using the




bioassay data.







     Both models are used to extrapolate from observed animal



tumor rates to expected human response rates.  The age-specific



human death rates due to liver cancers are used to estimate two



of the parameters in the human model.  The use of human data  for



this purpose  is mandated by two  factors.  Under the assumed



model, the agent acts on initiated, preneoplastic cells to



increase  their number.  As a result, the increase in the tumor



rate should be proportional to  the background number of preneo-



plastic cells, which in  turn is  proportional to the background



tumor rate.   This theoretical prediction is confirmed by the



strong dependence of the  sensitivity of  the tumor response  to the



background tumor rates observed in  five  separate animal bio-



assavs.   The  usual  assumption of the equivalent dose between



species being proportional to the cube root of  the  ratio of the



species weight is  also employed in  the development  of the human



dose response model.  Using  the developed  models it is



demonstrated  that  low-dose  linearity  can result from either form



of the  models under certain  circumstances.   In contrast,  is is



also shown that the log-logistic model with  a slope equal  to  3



gives prediction of risk that decrease  3 orders of  magnitude  for



 each order of magnitude  of reduction of  exposure.
-------
                              • - 4 -

      While both forms of the M-V-K model fit the observed data
 quite well, and both can be justified on theoretical and some
 experimental ground, their use for low-dose extrapolation leads
 to a wide variation in risk estimates.  For a lifetime daily dose
 of 0.1 ng/kg-day, or one order of magnitude below the animal
 experimental level, the estimates vary by a factor of more than
 10,000.

      Although the negative exponential form of the model results
 in a linear low-dose response relationship, the risk estimates
 derived  from it are two orders of magnitude below the upper bound
 values obtained by EPA when TCDD was  modeled as a complete carci-
 nogen.   This difference is primarily  due to the low human back-
 ground liver cancer rates  as compared to those of the female rat.
 Thus,  treating  TCDD as a promoter could  have a strong impact on
 any regulatory  decision.   However,  the model is based upon the
 assumption  that the site of cancer  induction in humans  would also
 be  in  the liver.   An analysis  of tissue  dose distribution and cell
 turnover rates  and receptor protein levels  for other organs or
 tissues should  be  conducted before  the equivalence  of target site
 between- humans  and rodents  is  accepted unequivocally.
2.
MATHEMATICAL DOSE RESPONSE MODEL FOR PROMOTERS
     A mathematical model has been developed by Moolgavkar,
Venzon and Knudson  (1979, 1981) that is based on a two-stage
model for carcinogenesis.  This model is depicted in Figure 1.
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-------
 According to the model,  the population of cells at risk is proli-

 ferating cells,  often referred to as stem cells.   Stem cells are

 those from which most other cells in an organ arise;  once a cell

 has  differentiated and left the pool of proliferating cells it  is

 no longer susceptible to heritable alterations of its DNA.   A

 normal,  susceptible stem cell may do one of several  things.  It

 may  divided into two daughter stem cells,  terminally differen-

 tiate, die,  or undergo mutation at a critical site that results

 in formation of  a preneoplastic or intermediate cell.   A preneo-

 plastic  cell has undergone  one of the changes necessary to  become

 a cancer cell but is not yet cancerous.   The cancerous cell will,

 after a  sufficient length of time,  divide  into enough cells that

 it becomes  a detectable  tumor.   All of these processes can  be

 described mathematically by postulating  specific  rates for  the

 cell changes.  Moolgavkar and  Knudson (1981)  showed  that to a close

 approximation, the age-specific tumor rate  at age  t  for their two-

 stage model  may  be expressed as follows:
          I(t) = MjjMj J"G0(v)|exp  [(B-D) (t-v) ]] dv
                      f"i      *•                 '
                                         (i:
where
     Kt)

     M,
      0
     M.
     C0(v)

     B
     D
age-specific cancer incidence at age t?

transition rate from stem to preneoplastic
cell;

transition rate from preneoplastic to
cancerous cell;

number of susceptible stem cells at age v;

birth rate or rate of cell proliferation of
preneoplastic cells; and

death rate of preneoplastic cells.
-------
                              - 6 -





     Essentially, the equation describes the progression from a



normal stem cell to a cancerous cell under the assumptions of the



two-stage model.  This model is a combination of deterministic



and stochastic components.  The numbers of preneoplastic and



fully malignant cells at any time are assumed to be random



variables that are dependent upon these event rates, while the



number of normal cells at risk of transformation at time t,



denoted by C0 (t) , is assumed to be deterministic and known.








     The biological processes described by the equation 1 can be



affected by the level of exposure to carcinogenic agents, and are



more likely to occur as the length of exposure increases; as a



result, they are exposure and time dependent.  Thorslund et al.,



(1987)  have described an exposure- and time-dependent version of



the model.  In the context of biological mechanisms of carcino-



genesis, the model can be used to predict the risk of agents that



exert their effects in a number of different ways.  Mutation-



inducing initiating agents could affect the transition rates



between cell stages (MQ or F^), while promoting agents may



increase the proliferation rate of preneoplastic cells  (increas-



ing B without affecting D) .  Cocarcinogens may increase the



proliferation rate of normal stem cells  (CQ), thereby increasing



the size of the target for initiating agents.  Inhibitors could



remove cells from the populations of susceptible stem or preneo-



plastic cells through toxicity  (e.g., increasing D without



affecting B) or by inducing differentiation.
-------
                              - 7 -





     For the purposes of the model, a mechanism of action for



TCDD has been postulated that is similar to that of a promoting



agent.  It has been assumed that TCDD exerts its carcinogenic



effect by increasing the preneoplastic cell birth rate.  The



difference between the birth and death rates is the net preneo-



plastic cell growth rate and is dose-dependent.  This rate is



denoted as:
          G(x)  =   B + B  (x) - D
                                                  (2)
where
G(x)  =   dose-dependent preneoplastic cell growth rate



B (x) =   the increase in the birth rate due to TCDD
          B,D       are as defined in equation  (1)



Upon maturity, the number of stem cells in the  liver may be



viewed as relatively constant.  Under this assumption, the term



C (v) may be taken to be equal to a non-time-dependent constant,



C .   Substituting these definitions for G(x) and C  (v) into
 o                                                o


equation  (1) and integrating yields:
                               G(x)t-l]/G(x)                 (3)



which is the age-specific tumor rate expressed  in a dose and time



dependent form.  The probability of a tumor by  time t  in the



absence of competing mortality is obtained from the well-known



relationship:
          P(x,t) =  1-exp- £ I(x,v)dv                        (4)



Substituting equation  (3) into equation  (4)  and  integrating



yields the expression:
-------
                              - 8 -





          P(x,t) =  1-exp- {M[expG(x)t-l-G(x) t]/G(x)2)       (5)








where M = M M-C  is a composite parameter that is proportional to



the background tumor rate.  The growth rate has a normal value



G(o) = B-D in the absence of exposure and has a finite upper



bound G () that is determined by how rapidly a cell can go



through its normal proliferative cycle.  Using this notation, the



exposure-dependent growth rate of preneoplastic cells may be



expressed as:
            G(x) =  G(o) + [G(«o)-G(o)]R(x)                    (6)








where R(x) is the fractional amount of the maximal  increase  in




the growth rate that can be induced by exposure at  level x.  The




function R(x) is dependent upon the specific mechanism by which



TCDD induces cell proliferation.  A specific form for R(x) can be



derived for each hypothesized mechanism of action.  The next




section discusses the factors that need to be considered in  the



selection of the R(x) to be used in equation  (6) .








3 .   MECHANISMS OF THE ACTION








     Carcinogenesis is a multistep process which displays at




least two distinct steps.  External agents  (carcinogens) can act



to augment the process at each of these steps  (Weinstein, 1981) .




For simplicity, carcinogens are often divided into  two classes:




"initiators"  (those that act at the first stage of  cancer devel-




opment) and "promoters"  (those that act at later stages and  hence
-------
                               -  9  -





 "promote"  the  action  of  initiators)  (Weisburger and Williams,



 1983; Weinstein,  1984) .   However,  recent  studies have shown that



 both initiation and promotion  are  more complex phenomena than is



 reflected  in this dichotomous  classification  (Becker, 1984;



 Upton, et  al., 1985;  Gallagher,  1986).  Promotion frequently



 involves the action of promoters on cell  membranes, including



 binding to protein kinases,  inhibition of intercellular communi-



 cation, and other effects on membrane structure and function, but



 also may involve genetic  changes,  including methylation of DNA,



 modulation of  expression  of  genes  involved in cell differentia-



 tion, activation of oncogenes, and chromosome damage  (Shank,



 1984; Yamasaki and Weinstein,  1985; Weinstein, 1984; Jones, 1986),
     It is unlikely that TCDD is a typical initiator.  There is



little evidence that it causes point mutations in bacteria  (e.g.



in the Ames test) but it has some genotoxic potential, as



reflected by a clear positive result in the mouse lymphoma assay



(Rogers, et al., 1982).  TCDD has little propensity for covalent



binding to DNA (Poland and Glover, 1979); however, it is known



that after binding of TCDD to a cytosol "receptor" protein, the



TCDD-receptor complex is translocated to the nucleus of the cell



where it interacts with different sites on DNA and can affect



various functions.  These include regulating transcription of the



cytochrome P-448 gene (Israel and Whitlock, 1984; Whitlock, et



al., 1984; Jones, et al., 1985); regulating the levels of other



enzymes such as ornithine decarboxylase, DT-diaphorase, and UDP-



glucuronyl transferase; and affecting the regulation of cell



proliferation and differentiation by interfering with the ability
-------
                              - 10 -






of other sites on DNA to bind receptor complexes such as those of



epidermal growth factor and glucocorticoids (Poland and Knutson




1982) .







     The latter two functions may be those by which TCDD affects



tumor promotion.  Agents that increase the number of proliferat-



ing cells by increasing cell proliferation rates or decreasing



the number of cells that terminally differentiate could increase



the likelihood of mutational events by tumor initiators or permit



clonal expansion of initiated cells.  The mechanism of enhanced



cell proliferation may involve the interaction of a single TCDD-



receptor protein complex with a single site on DNA  (such as that



of a regulator gene), which would be consistent with a fractional



change in a growth rate function of the form R(x) = 1-exp-Vx.



Alternatively, TCDD's action may require the formation of



multiple complexes with roles at multiple sites on DNA, which



would suggest a log-logistic relationship for



               R(x) =  [l+exp-(I+S In x)]'1.








     Studies performed to date evaluating the ability of TCDD  to



cause hepatic cell proliferation are inconclusive.  Conaway and



Matsumura  (1975) ,  for example, found an increase  in hepatic



nuclear  3H-thymidine activity as compared to controls of 56% or



94%, depending on  the method of preparation, ten  days after adminis-



tration  of a  single oral dose of 5 ug/kg TCDD to  male Sprague-



Dawley rats.  Dickens et al.  (1981)  saw an  increase of  94% in



3H-thymidine  activity  associated with  hepatic DNA five  days  after
-------
                               -  11  -


a  similar dose  although  the difference was not  statistically

significant due to  the small  number of animals  used.  These authors

also noted significant increases  in liver net weight and relative

liver weight compared to body weight  in treated animals.  In a

similar experiment, however,  Christian and Peterson  (1983) saw no

such increases.  Further studies  are  clearly needed to resolve

the discrepancies along  with  studies  of the dose-response and

time-course relationships of  any  increases observed.



     In the absence of detailed,  reliable data  on the dose-

related effects  of TCDD  on cell growth rates, we shall assume two

functional forms for R(x):



    (1)   R(x)  =  1-exp-Vx     which  would be appropriate if the

proliferation effect was due  to the interaction of a single TCDD-

receptor protein complex with a single site on  DNA, or
     (2)   R(x) =  [l+exp-(I+S In x)]      which would be appropriate

if the interaction of multiple TCDD-protein receptor complexes at

multiple sites were required for stimulation of the growth rate.


4.   RATIONALE FOR THE SELECTION OF THE DATA USED TO DEVELOP
     TUMOR DOSE RESPONSE MODEL


     A number of decisions regarding the most appropriate study,

tumor end point, time frame, pathological diagnosis, and exposure

parameter must be made in order to obtain the specific data set

that is used to estimate the parameters in the postulated dose

response models.  The rationale for those decisions is discussed

in this section.
-------
                              - 12 -





4.1  Selection of Animal Tumor Data








     As discussed in the Health Effects Assessment Document for



Polychlorinated Dibenzo-p-Dioxins (USEPA, 1985),  several chronic



animal studies of TCDD have demonstrated enhanced liver tumor



rates.  In its extrapolation from animal data to human cancer



risk estimates, EPA used the hepatic tumor rates in female



Sprague-Dawley rats from the Dow 2-year feeding bioassay (Kociba



et al., 1978).  These data are discussed in some detail in both



the HAD and in the issue paper on the quantitative implication of



the use of different models.  Every U.S. and foreign agency



except one (California) has used the DOW study as its pertinent



cancer study for extrapolating to humans.  In this report, only



the liver response will be modeled.  While the female rats in



this study also developed rare tumors of the hard palate/nasal



turbinates as well as infrequent lung tumors, neither of these



tumor types has been shown to be the result of promotion.  On the



other hand, TCDD has been shown to be a potent promoter in the



rat liver  (Pitot et al., 1980).








     In its extrapolation procedure for TCDD, EPA uses the



pathology analysis results of two independent pathologists, Dr.



R. Kociba and Dr. R. Squire.  Even though Squire's analysis



included more liver tumors in the control, low- and mid-dose



groups, the estimates of incremental increase in cancer risk



differed by less than  10%.  This report will use the Squire



pathology analysis of the data.  Although the higher background
-------
                              -  13 -





rates based on the Squire analysis lead to higher incremental



cancer risks for animals under a promoter model, the human back-



ground rates are substituted when the results are extrapolated to



humans.  The final estimates derived for humans should be



similar, therefore, since they depend only on differences in



response between the animal dose groups, which were about the



same according to both pathologists.








     In an attempt to adjust for high early non-tumor related



mortality in the high dose group, EPA eliminated from considera-



tion all animals dying prior to the 13th month, when the first



liver tumor was observed.  This censored data set also will be



used here.








4.2  Selection of Exposure Variable
     Rose et al.  (1976) applied a simple first-order  (i.e., one



compartment open) model to estimate the steady-state concentra-



tions of TCDD in the liver of female Sprague-Dawley Spartan rats.




Based upon their analysis, it was concluded that the administered



dose following oral exposure was proportional to the target organ




dose (i.e., steady-state liver dose) in the oral dose range of



0.01-1.0 ug/kg/day TCDD; however, actual steady-state doses were



not measured.  Kociba et al.  (1978) measured the levels of TCDD




at the end of a two-year carcinogenesis bioassay in five rats



from each exposure group receiving 0.001, 0.001 or 0.1 ug/kg/day.



Levels of 0.54, 5.1, or 24 parts per billion, respectively, were




observed.  Portier et al.  (1984) used these exposure levels to
-------
                               -  14  -



 fit dose-response models to  the  tumor data.  To extrapolate expo-


 sure  levels  from those received  in  the bioassay to the lower levels


 encountered  environmentally, the relationship between administered


 and target organ dose was assumed to be  linear from 0.01 ug/kg/day


 to zero.  This approach has  several problems.  The results are


 based on small sample sizes  with considerable variability between


 measurements.  Potential biases  may also have been introduced by


 obtaining observations only  at terminal  sacrifice and by having


 high tumor rates and liver weight increases at the highest dose


 level, which distort the levels  of TCDD  per cell'when measured as


 proportional to liver weight.  As a result of these problems, the


 average liver exposure level data will not be used to .derive dose-


 response models.  Nevertheless,  since the value of the highest dose


 used in the bioassay is the  only one that would be altered by using


 these data, it can be shown  that they have virtually no influence


on the estimate that results from the models employed in the


 subsequent analysis.  Furthermore, the use of administered doses


 is consistent with EPA's previous approaches so that direct


 comparisons are more meaningful.




 5.  PARAMETER ESTIMATION





     The two models that will be used to predict risk may be


expressed as:
     P(x,t) = 1-exp-MJ[expG(x)t-l-G(x)t]/G(x)2 I              (8)


       G(x) = G(o) + [G(«o)-G(o) ]R(x)


       _, v _ l~exp-Vx                                 case  1
        ' ' ~                   _ i
              [H-exp-d+S In x) ]                       case  2
-------
                               - 15  -





The  number  of  parameters  to be estimated is  4  (i.e.,  M,  G(o),



G(^, V)  for the  negative exponential  preneoplastic  cell growth



model  (case 1) , and  5  (i.e.,  M,  G(o),  G («o) ,  I,  S)  for the log-



logistic  preneoplastic  cell growth  model (case  2).   This section



explains  how the  parameters were estimated and  presents  the  results



obtained  for the  most informative dose response relationship.
     As discussed  in  the  previous  section,  the most  reliable and



biologically meaningful data  set that  can be  used  to obtain a



dose response relationship  for  TCDD  is liver  tumors  in  female



rats surviving one year in  the  DOW study using the Squire



pathology analysis.   This data  set is  shown in Table 1.








     Of the four or five  unknown parameters in the models, M and



G(o) do not depend on exposure.  The parameter M is  proportional



to the product of  the background cell  transition rates  and G(o)



is proportional to the preneoplastic cell growth rates.  If



time-to-tumor data were available, these parameters  could be



estimated from control data.  When x=o the  exposure-time model



has the form:








          P(o,t) = l-exp[-MG(t) ]                             (9)
where
G(t)  = [exp(G(o)t)-l-G(o)
In the absence of reliable time-to-tumor data, G(o) may be



specified based upon knowledge of cell turnover rates or time
-------
                              - 16 -




dependent tumor occurrence.  The approach taken here was to find


a range for G(o) that is consistent with the age-specific tumor


rate increasing from the 3rd to the 5th power of age.  This range


was found by Cook et al. (1969) to be consistent with the tumor


registry data for most tumor sites in eleven different popula-


tions.  Using the multistage model, the probability of a tumor


occurrence by time t may be expressed as
          P(o,t) = l-exp-Atk       where 4^k^6              (10)
An estimate of A can be obtained by fixing k and and substituting


a background rate estimate at a fixed time in equation  (10),


yielding
          -ln[l-P(o,t)]/tk = A                              (11)
The background rate is obtained  from  the vehicle  control, which



results in an estimate of






          P(o,t) = 16/85 =  0.1882






Substituting t=104, P (o, t) =0 . 1882 ,  and  k=4  or  6  into  equation

                                 -9                        -13
 (11) gives the results A=l. 78x10   when k=4 and  A=l. 65x10    when


k=6.  To estimate G(o), we  transform  equation  (9)  to  the  form
                        =  ln(M/G(o)2)+ln[expG(o)t-l-G(o)t]   (12)
-------
                              - 17 -




which for G(o)t^3  is approximated closely by the simple rela-


tionship
                           = ln(M/G(o))+G(o) t
(13)
This equation is then equated to numerical values obtained from


equation  (10) at t=52,104, which gives two linear equations and

                                                                2
two unknowns for each k that can be used to solve for  ln(M/G(o)  )


and G(o).  These values are used in turn to estimate M.  Follow-


ing this approach the values shown in Table 2 were obtained.  The


purpose of this manipulation is simply to obtain a time-to-tumor


relationship for the M-K-V model that corresponds to that which


has been previously observed often in terms of the multistage model,


Taking the range of the parameters is done in order to investi-


gate the sensitivity of the assumed parameters in the final risk


estimates.  The observed tumor data are not used to estimate G(o)


so that a valid goodness-of-fit test can be obtained from the


four data points, which is an added advantage of the approach.


The remaining two parameters, G («o) and V, are obtained by equating


the parametric form of the model to the observed rates at the two


highest doses and solving the resulting simultaneous non-linear


equations for two unknowns.  This is done for both values of k


but since the resulting models give virtually identical final


results, only the values for k=4 are shown in Table 3, along with


the corresponding expected values under the model.





     The log-logistic cell growth model is also  fitted to the


data.  This is done by assuming the slope is equal to  1,2,  or 3
-------
                              - 18 -


common values in many biological systems and estimating the

intercept corresponding to each assumed slope value.  As can be

seen in Tables 1 and 3, values of the parameters that provide the

best log-logistic fit to the data are S=3 and 1=14.5.  The

results of this analysis are displayed in Table 3.  The estimates

for G (°o) and I were obtained from the tumor response data using

the same methodology as was done for the negative exponential

model.



     In the next section the validity of the fitted models is

tested in a variety of different ways.



6.   EVIDENCE FOR THE VALIDITY OF THE OBTAINED DOSE RESPONSE
     MODELS


     The validity of the dose response models that were obtained

in the previous section can be evaluated in a number of ways.

Since the models are a subset of the M-V-K model, which has been

shown to have a remarkable ability to describe a variety of

different carcinogenic phenomena and age-specific cancer rates in

humans, a degree of acceptance for their application to TCDD

should be accorded on purely theoretical grounds.



     In addition, the predictions of the model can be evaluated

using  (1) the goodness-of-fit of the model to the tumor data from

which it was derived,  (2) its consistency with TCDD-induced

proliferation data from separate experiments,  (3) its ability  to

predict the TCDD-induced tumor responses in other sexes and
-------
                               -  19  -



 strains  of the same species  (rats)  using  adjustments  only  for


 background tumor rates,  (4)  its  ability to  predict  the TCDD-


 induced  tumor rates in  another species  (mice), making adjustments


 for  background rates and  estimating a separate growth rate para-


 meter  for  that species, and  (5)  its consistency with  the predic-


 tion that  the slope of  the linear relationship between the log of


 the  age-specific tumor  rate  regressed against age will be a


 raonotonically increasing  function of exposure.




     In  this  section, the models are evaluated using  the first


 four of  these criteria.   The consistency  of the dose-dependent


 slope  (i.e.,  criterion  (5))  was not attempted due to  time and


 resource constraints.




 6.1  Goodness-of-Fit of Models to DOW Female Rat Data
     The goodness-of-fits of the hypothesized models to the data


set from which they were derived are shown in Table 1.  As is

                  2
indicated by the X  values, both models  fit the data adequatelv.


A better fit could have been obtained by letting the background


rate parameter deviate from the observed background; however,


considerable attention has been paid to  the fact that the low-


dose tumor response is lower than the control tumor response,


which has been suggested to imply that some type of unspecified


compensatory low-dose mechanism is operating  (see Sielken, 1987).


The goodness-of-fit test for the model fitted using only three


data points has a more specific meaning  and greater power than
-------
                              - 20 -





using all the exposure levels to obtain parameter estimates.  The



null hypothesis in this case is that the low-dose information is



consistent with our hypothesized tumor dose response model.  As



indicated previously, there is no evidence that is inconsistent



with this hypothesis, which is a stronger criterion than a



general goodness-of-fit model.  The better fit of the log-



logistic model compared to the negative exponential model should



not be interpreted as suggesting that the former model is more



valid, since both models fit the-data adequately.  Equivalently,



the slight improvement in fit due to increasing the slope in the



logistic model should not be viewed as strong evidence for a



large slope.  We note in Table 3 that the log-logistic model fits



the data adequately for slope values from 1 to 3, even though, as



is indicated by the estimated rat cancer risks, the implications



for low dose extrapolation are very different.  To illustrate the



type of results one might obtain with a log-logistic model, we



use the case with a slope of 3 in the final evaluation in addition



to the negative-exponential model.







6.2  Consistency with Cell Growth Rate Data








     The form of R(x) (i.e., dose-dependent change in growth



rate) is obtained from theoretical considerations and the shape



of the liver tumor dose-response model.  Ideally, the growth



rates of hepatocellular adenomas would be used to estimate the



parameters in R(x).  The hepatocellular adenomas induced by TCDD



are thought to be a preneoplastic stage of hepatocellular
-------
                               -  21  -






carcinomas.  A more  practical  but  less  direct  alternative would



be to measure the  TCDD-induced growth or  turnover  rates  in normal



liver cells  in the same  species  for which TCDD-induced tumor dose



response  information exists.   One measure that can be used in



this regard  is the H  thymidine  levels  incorporated  in liver cell



nuclei  after TCDD  exposure.  As  cited by  EPA  (1985), page 8-20,



it was  shown that  the H-thymidine  activity at control and a 5



ug/kg TCDD exposure  levels were  29  and  45 cpm/mg liver,



respectively.  We  assume that  the growth  rate  is proportional to



this index and that  the maximum  growth  rate is obtained  at the



5 ug/kg exposure level.  If multiple exposures had been  used, it



might have been possible to define  the  shape of R(x), as is



indicated in Figure  1.  However, it is  possible to obtain an



estimate of  G( )  from the limited data  available.  The ratio of



the maximum  to minimum growth  rates (G( )/G(o)) can be estimated



by taking the ratio  of the H   thymidine counts.  Multiplying this



ratio (45/29) by the  assumed G(o)=0.0533  gives the value G( ) =



(45/29)xO.0533=0.0827, which compares closely  with the value of



0.0817 obtained from  the shape of the bioassay curve.








6.3  Consistency of Model With Other Rat  Dose-Response Data
     Under the assumed M-V-K model, a promoting agent acts on




initiated, preneoplas'tic cells to increase their number.  As a



result, the increase in tumor rate should be proportional to the




background number of preneoplastic cells, which in turn is



proportional to the background tumor rate.  The background tumor
-------
                             - - 22 -






rate parameter M should be different, therefore, for strains and




sexes that have varying background liver tumor rates.  To test




the validity of the derived model, all parameters except M were




assumed to be known and the resulting model was fitted to the NTP




bioassay data for male and female Osborne-Mendel rats.  The



results of this analysis are shown in Tables 4 and 5 for female




and male rats respectively.  We note that the resulting model




gives an adequate fit to both data sets even though the non-




background tumor rates have little weight in the parameter




estimation.








6.4 Consistency of Model With Mice Dose-Response Data








     Due to biological differences, we would expect that enzyme .



and receptors levels would vary between species.  As a result, in



addition to background levels, a parameter for the exposure-




induced change in the growth rate V would also have to be



estimated when using a different species.  For the male mice data




in the NTP study, two parameters  (M and V) were estimated and the




other parameters taken from the previous model.  For the female




data, only one parameter  (M) was estimated while V was obtained



from the male data.  The results of this analysis are shown in



Tables 6 and 7 for male and female mice, respectively.  We note



that the resulting models also are consistent with the observed




data.
-------
                               - 23 -
 7.
HUMAN DOSE RESPONSE MODEL
      In this final section,  human age-specific cancer death rate
 data and parameters estimated from the animal bioassays  are
 combined to obtain a human dose-response model.   The rationale
 for this approach is that the background tumor rate  parameter is
 the critical factor in determining species  sensitivity.   This
 rationale is supported by the results  in Table 8,  which  demon-
 strate  the consistency of risks  among  species after  adjusting for
 background rates.

      The human  age-specific  liver cancer death rates can be used
 to  obtain estimates of the human parameters M and  G(o) under the
 assumptions  that  most liver  cancers  are fatal and  that these
 rates represent actual background rates.  Under  the  model,  the
 age-specific background tumor rates  may be  expressed as
               =M[exp(G(o) t)-
                                                      (14)
     Since we expect G(o)t>3  for most  of  the  observed  age range,
taking the natural log of eq.  14 yields
                  £Jln  [M/G(o)]+G(o)t
                                                      (15)
Values for I(t) are derived from Table  1-25, Section  1, Vol. II,
Part A, of the Vital Statistics of the  United States  1980  (U.S.
HHS, 1985) and the 1980 U.S. Census.
-------
                              - 24 -
     Using these data, a simple linear regression of the form



In[I(t)]=A+Bt yields a correlation coefficient r=0.99 with values



A=-15.6870 and B=0.09381(years)~   (see table 9).  M and G(o) are



then estimated by setting the linear regression equation equal to



eq.(15).  Since A and B are now known, the intercept A=ln[M/G(o)]



and B=G(o).  Estimates of the parameters are thus
     M=B(expA)=1.44xlO~8 and G(o)=.09381  (years)'1          (16!
     The maximum growth rate G (<*9 is assumed to be the same per-



cent change over background as was observed in the DOW female rat



data.  Thus, for the human model G(«°) =0 . 09381x (0 . 0817/0 . 0533) =0 .1438 .



Finally, the parameters V and I for humans are obtained from the



DOW female rat study by using the usual surface area correction



(70/0.45)   =5.38.  The value of the human parameters so obtained



are shown in Table 10.  The resulting mathematical extrapolation



model using these parameters and risks  at various low environ-



mental exposure levels are shown in Table 11.
-------
  Table  1 -  Comparison  of  Observed  and  Predicted  Liver  Tumor
              Rates  in the  DOW TCDD  Feeding  Study  Using  Female Rats
Exposure
ug/Kg/day
(X)
0.0
0.001
0.01
0.1
Number of
Animals
Surviving
First Year
85
48
48
40
i • 4

Number
Observed
16 (19%)
8 (16%)
27 (56%)
33 (82%)
2
X
d.f.
P value
(%) of Animals With Liver Tumors |
Predicted
1st order Equilibrium
16.0 (19%)
10.8 (23%)
27.0 (56%)
33.0 (82%)
0.96
1
0.32
Log-Logistic S»3
16.0 (19%)
9.4 (20%)
27.0 (56%)
33.0 (82%)
0.16
1
0.78
Parameters
Symbols

G(o)
G(<~)
M
V
I
S
Estimates
1st order] Log-Logistics
0.0533
0.0817 ,
2.3798x 10
109.51
14.50
3.00

Values Derived From
See text
Estimated From Data
Estimated From Data
Estimated from Data
Estimated from Data
Assumed
    P(xft) - 1-exp-M  [expG(x)t-l-G(x)t]/G  (x)  ,  t=104

      G(x) * G(o) + [G(o*)-G
-------
Table 2 - Parameter Estimates For The Moolgavkar-Knudson
            Model That Simulate Range of Multistage Time
            to Tumor Model for Usual Range of Human Data
Assumed Value
of Time
Parameter
4
6
1980 U.S.
Vital Statistics
ICD 155.
Liver Cancer
Deaths
Resulting Parameter
Estimates
G
0.0533
0.0800
0.0631a
M
2.15xlO~6
3.03xlO~7
1.44xlO~8
     Value adjusted by .0938(years)~ x(70 years/104 weeks)
-------
Table 3 - Effects of Varying Slope Parameter in Log-
          Logistic Model on Goodness-of-Fit and Low-Dose
          Risk Estimates Based on DOW Study
Parameter
Svmbol
M
G(o)
G(°o)
I
S
Augmented
Rat Risk at
x=lxlO~3
x=lxlO~4
x2
P
Parameter Estimate
2.3798 x 10~6
0.0533
0.0817
5.29
1
6.3xlO~2
6.6xlO~3
2.066
0.18
9.90
2
6.6xlO~3
6.6xlO~5
0.245
0.62
14.50
3
6.6xlO~4
6.6xlO~7
0.155
0.72

oo
0
0
0.146
0.73
I
I
-------
Table 4 -. Comparison of Observed and Predicted TCDD-Induced
           Liver Tumor Rates  In the Osborne-Mendel Female
           Rats From the  NCI  Gavage Study..
Exposure
ug/kg/day
(X)
Historical
Control
0
0.0014
0.0071
0.0714
Number of
Animals
Exposed

970
75
49
50
49



Number (%)
Observed

21 (2.2)
5 (6.7)
1 (2.0)
3 (6.0)
14 (28.6)
2
X
d.f.
P value
of Animals with Liver Tumors
Predicted

2.8 (3.7)
2.4 (4.9)
5.4 (10.8)
13.2 (26.9)
3.99
3
0-.28
Values of Parameters (see text)
Symbols
G(o)
G(o«)
V
M
Biological
Interpretation Estimates
background cell
growth rate 0.0533
maximum cell
growth rate 0.0817
incremental rate
change per
unit dose 109.51
background _
mutation rate 4.3036 x 10~
Values drived from
Assumed (A priori)
Dow Study
Dow Study
Estimated From
Data
 P(x) -  1-exp-M •[ [expG(x) t-l-G(x)t] /G

 G(x) =  G(o')  + [G(<*>)-G(o) ] [1-exp-Vx]
-------
Table 5 - Comparison  of  Observed and Predicted TCDD-Induced
           Liver Tumor Rates  in  Osborne-Mendel Male Rats from
           the NCI Gavage  Study.
Exposure
ug/kg/day
(X)
Historical
Control
0
0.0014
0.0071
0.0714
Number of
Animals
Exposed

975
74
50
50
50

Number (%)
Observed

9 (0.9)
0 (0.0)
0 (0.0)
0 (0.0)
3 (6.0)
x~
d.f.
P value
of Animals With Liver Tumors
Predicted

0.3 ( .4)
0.2 ( .5)
0.6 (1.2)
1.6 (3.2)
2.37
3
0.58
Values of Parameters (see text)
Symbols
G(o)
G(oo)
V
M
Biological
Interpretation
background cell
growth rate
maximum cell
growth rate
incremental rate
change per
unit dose
background
mutation rate
Estimates
0.0533
0.0817
109.51
4.420 x 10~8
Values drived from
Assumed (A priori)
Dow Study
Dow Study
Estimated
From Data
P(x)  =  1-exp-M |[expG(x)t-l-G(x)t]/G2(x)]


G(x)  =  G(o)  + fG( «=>)-G(o) ] [1-exp-Vx]
-------
Table 6 - Comparison of Observed and Predicted  TCDD-Induced
           Liver Tumor Rates  in B6C3F1 Male Mice  from  the  NCI
           Gavage Study.
Exposure
ug/kg/day
(X)
0
0.00143
0.00714
0.07143
Number of
Animals
Exposed
73
49
49
50



Number (%) o
Observed
15 (20.5)
12 (24.5)
13 (26.5)
27 (54.0)
2
X
d.f.
P value
f Animals With Liver Tumors
Predicted
15.00 (20.5)
10.40 (21.2)
11.77 (24.0)
27.77 (55.5)
0.53
2
0.77
Values of Parameters (see text)
Symbols
G(o)
G(«o)
V
M
Biological
Interpretation
background cell
growth rate
maximum cell
growth rate
incremental rate
change per
unit dose
background
mutation rate
Estimates
0.0533
0.0817
13.19
2.6248 x 10~7
Values drived from
Assumed (A priori)
Dow Study
Estimated
From Data
Estimated '
From Data
 P(x) = 1-exp-M

 G(x) = G(o)
>-M  | [expG(x)t-l-G(x)t]/G2(x) |

+ . [G(oo) -G(o) ] [1-exp-Vxl
-------
Table 7 - Comparison of Observed  and Predicted TCDD-Induced
           Liver Tumor Rates  in B6C3F1  Female Mice from the
           NCI Gavage Study.
Exposure
ug/kg/day
(X)
0
0.006
0.028
0.286
Number of
Animals
Exposed
73
50
48
47



Number (%) c
Observed
3 ( 4.1)
6 (12.0)
6 (12.5)
11 (23.4) .
2
X
d.f .
P value
if Animals With Liver Tumors
Predicted
.3.7 ( 5.1)
2.9 ( 5.9)
4.5 ( 9.4)
16.0 (34.0)
6.46
3
0.09
Values of Parameters (see text)
Symbols
G(o)
G(«<>)
V
M
Biological
Interpretation
background cell
growth rate
maximum cell
growth rate
incremental rate
change per
unit dose
background
mutation rate
Estimates
0.0533
0.0817
13.19
5.9884 x 10~7
Values drived from
Assumed (A priori)
Dow Study
NCI-Male Mice
Estimated
From Data
P(x) =  1-exp-M | [expG(x)t-l-G(x)t]/G2(x)

G(x) =  G(o)  + [GC*3 )-G(o) ] [1-exp-Vx]
-------
 IQ
 to

 en
 >,
 m
 (0
 (0
 (0
 o
•H  cn
PQ  Q)
    -P
 C  (0
 IU «
 OJ
CQ  O
    M
 to  tn
^  X
 w  o
•H  (0
cc;  CQ

u-4  m
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 o o
 (0 -H
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 M O
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 Es fe
 o
oo

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rH
JQ
 It
EH


































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T3
n
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M
60

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PJ


























w
•H
Pi

3
O
M
61
O
cd






01
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C
3
O
M
60

O
Cd
«
<



u
cd



























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,
0)
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cd
o
i
<
OO
cr.
i — i

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1
1


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1
0
— 1
X
1 — 1
o
CM
"d"
•
"*







o
•
o
") ~;

0
i — i
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CU
2
I

m
ON
o
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r ^

0
^-<
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1 — 1
fa
CO
u
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01
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e
CU
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o
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oo
~-<


e»s
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sf
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-------
                       TABLE 9

    Estimation of Background Human Lifetime Liver Cancer
    Rate  (M) and Background Human Liver Cell Proliferation
    Rates G(o) by Fitting In[I(t)]=A+Bt=ln(M/G(o))+G(o)t
                    1980 U.S.
                  Liver Cancer
1980 U.S.
Age
25-29
30-34
35-39
40-44
45-49
50-54
55-59
60-64
65-74
75+
(Mid-point)
27.5
32.5
37.5
42.5
47.5
52.5
57.5
62.5
70
80
Deaths
ICD (155)a
36
55
53
95
173
289
522
699
1795
1805
Population
(thousands)
19,518
17,558
13,963
11,668
11,088
11,709
11,614
10,086
15,578
9,973
Rates/
100,000
0.1844
0.3132
0.3726
0.8142
1.5602
2.4682
4.4946
6.9304
11.5227
18.0970
All races, both sexes
    In I(t)=-15.687+0.098381t
-------
Table 10 -  Parameters  in Human Dose Response Model
                Parameters
Symbols

G(o)
G(oo)
M
V
I
S
Estimates
1st order I Log-Logistics
0.0938
0.1438
1.44 x 10"8
589.16



19.55
3.00
Values Derived From
Human Mortality Data
Proportional to DOW Female
Rats
Human Mortality Data
Surface Area x DOW Female
DOW Female + 3 x Log
Surface Area
Assumed
-------
     Table 11 - Estimates of Low-Dose Incremental Ri
               Exposed to 2,3,7,8-TCDD a'D'C'Q°
                                                  Risk To Humans
                               MODEL
Promotion
Dose
ng/kg-day
io-5
1C'4
io-3
1(T2
io-1
1
Negative
Exponential
1.7 x 10~8
1.7 x IO"7
1.7 x 10"6
1.7 x 10~"5
1.8 x 10~4
2.4 x 10"~3
Log-
Logistic
_—
—
-------
 P60
T©  "TCOD
                                               TV*io*s
     1O4
                        8
 .1
          M
0.0,
                                                              10
-------
2.—
                                               Oft
                                                   M/CC
                     I - exp - M



               wlnere  -c » 104
                    -I -G(M
PROBABIUT/OF
  0*01
                       s-.f? x  10
        *7    ivrp
                                        -2-
-------
Ho:
(|)
   a

(2)
                         3—  Ofos«rved *n<£  Postulated

                           H3"Tfi«f»*>«J'l>ve Co u 10^5. in,  Rat*
                           E-gposove "ha  ~T"GPD.,n  ^ :

                        count" is  Surro^«de -fov-
                                    1    °
                                -fr*e.t/o«*l
                    Cells  is
                  |4
          H
     in nuclei rat
                                                                 TCPD
                                                      - 1-533
-------
                    Figure 4 "•
         of
        of
.7
.3

 .2

 .I
    0.0   .001 .004 .006 .008.0(0 ,OIZ •0(4*0(6  -O(8
-------
U-S-  L
                                       ISS  and
2S  30   35
40  45   50  55
                                               *80    fy *
-------
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     P.-450, Gene Expression.  In Banbury Report 18.  Biological
     Mechanisms of Dioxin Action.  A. Poland, and R.D. Kimbrough,
     editors.  Cold Spring Harbour, N.Y.  1984.

Yamasaki, J., and I.E. Weinstein.  Cellular and Molecular
     Mechanisms of Tumor Promotion and Their Implications for
     Risk Assessment.  In Methods for Estimating Risk of Chemical
     Injury: Human and Non-Human Biota and Ecosystems, V.B. Vouk,
     G.G. Butler, D.G. Hoel, and D.B. Peakall, editors.  John
     Wiley and Sons, New York.  1985.
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                                                 March 1988
                                                 Review Draft
                 APPENDIX B
      EPIDEMIOLOGIC  CANCER  STUDIES  ON

     POLYCHLORINATED DIBENZO-£-DIOXINS,

          PARTICULARLY  2,3,7,8-TCDD
               David  L.  Bayliss
         Carcinogen Assessment Group
Office of Health and Environmental Assessment
      Office of Research and Development
     U.S. Environmental Protection Agency
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                                  INTRODUCTION

     Since the publication of the Health Assessment Document  (HAD)  for Poly-
chlorinated Dibenzo-p-Dioxins (U.S.  EPA, 1985),  many new  epidemiologic studies
have been reported in the literature dealing with the risk  of soft  tissue
sarcoma (STS) and/or non-Hodgkin's lymphoma in  users and  producers  of phenoxy-
acetic acid herbicides and chlorophenols.   Furthermore, the major positive
studies have come under intense  criticism.   The earlier HAD document concluded
that there was limited epidemiologic data regarding the carcinogenicity  of
polychlorinated dibenzo-p-dioxin-contaminated phenoxyacetic acid herbicides
and/or chlorophenols based mainly on the Swedish case control  studies, but the
evidence regarding 2,3,7,8-tetrachloro-dibenzo-p_-dioxin  (2,3,7,8-TCDD) was
judged inadequate due to the inability ,of any of the data up  to that time to
demonstrate that the risk was due to 2,3,7,8-TCDD alone and not to  one or more
of the other 74 isomers of polychlorinated dibenzo-_p_-dioxins  found  in the
phenoxyacetic acid herbicides and/or chlorophenols or the phenoxyacetic  acid
herbicides or chlorophenols free of 2,3,7,8-TCDD.
     This report is a review and brief analysis of the epidemiologic evidence
to date.                                           .    .    .         .

                              SOFT TISSUE SARCOMA

     The main studies supporting the finding of an excess risk of STS were the
two independent Swedish case control studies of Harden and Sandstrom  (1979) and
Eriksson et al. (1979, 1981). These investigators reported statisticially sig-
nificant (five- to sevenfold) elevated risks of STS from  occupational exposure
to phenoxyacetic acid herbicides and/or chlorophenols either  alone  or separately,
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some of which are known to contain 2,3,7,8-TCDD as an  impurity.   Eriksson et
                           '
al. further subdivided their cases and controls into two categories  based on
expected presence or absence of 2,3,7,8-TCDD among the phenoxyacetic acids
after removing from the group those persons that had been exposed to chlorophenol
and then recalculating the risk of STS.   The relative  risks  were  17  in  the
group exposed to 2,3,7,8-TCDD-contaminated phenoxyacetic acids [2,4,5-trichloro-
phenoxyacetic acid (2,4,5-T)] versus 4.2 in the group  exposed to  other  phenoxy-
acetic acid herbicides [(2,4-dichlorophenoxyacetic acid (2,4-D)  and  4-chloro-
2-methyl-phenoxyacetic acid (MCPA)] not believed to contain  2,3,7,8-TCDD.  This
seemed to imply that either 2,4,5-T appears to be associated with a  high risk
of STS or the polychlorinated dibenzo-p_-dioxin contaminants  (such as 2,3,7,8-
TCDD) within are contributing to the high risk of STS.
     The Swedish studies have been severely criticized in the scientific liter-
ature because of known or alleged methodology flaws and biases.   These  criticisms,
which are outlined in a report provided by the Australian Royal  Commission  (1985)
on the Use and Effects of Chemical Agents on Australian Personnel in Vietnam
include the following:  recall bias, unreliability of the exposure data, infor-
mation bias, observation bias, absence of a significant risk at any  single
specific cancer site, significant confounding factors, and lack of support  from
other studies.  (Although not specifically mentioned in the report,  the Eriksson
et al. case-control study is included in this critique by inference.)
     Some of the observations of the Australian Royal  Commission have validity.
Recall bias certainly may have been present.  Persons who have been  diagnosed
with a health problem hypothesized to be associated with exposure to a  given
agent such as a phenoxyacetic acid herbicide are more likely to be reminded of
a possible link with  that herbicide than is someone diagnosed with a different
health problem not hypothesized to be connected with that agent.   Hence, they
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are more likely to "recall" that exposure when questioned  about  it.  However,
given the rather exceptionally high risk ratios found by both  investigators in
their separate studies, it does not seem likely that recall  bias is the explana-
tion.  At the risk ratios calculated the power exceeds 99% in  both the Harden
study as well as the Eriksson study.  Even if the positive responses given by
the authors to the possibility of exposure were incorrect  on 33% and were
discarded, these studies would still retain 80% power i.e.,  confidence that the
recalculated significant Odds Ratio (OR) measures a true association.  Recall
bias was discussed somewhat in the HAD for polychlorinated dibenzo-p-dioxins
(U.S. EPA, |1985).
     Also, !there is a lack of substantiation of quantity of exposure to the
phenoxyacetic acid herbicides as well as to the chlorophenols.   Harden described,
in his doctoral dissertation dealing with the same cases as in his case-control
study, which herbicides andichlorophenols his cases and controls were exposed  to,
for how long, and when they were exposed; actual  quantities were not provided
although latency was determined.  Unfortunately,  neither investigator saw a need
to provide|a differential analysis by latency or by quantity or  length of exposure
to the herbicides and/or chlorophenols.   But, of course, at the  time, the authors
probably did not think such an analysis would be important.  They did not have
the vision of hindsight.  Harden and Sandstrom (1979) and Eriksson;et al.
(1979, 1981) contacted employers to verify exposures reported  in responses to
questionnaires sent to study members and survivors.  The response from employers
regarding use of phenoxyacetic acid herbicides was, "uncertain  and difficult to
evaluate" due to poor record keeping by the employer.  But for chlorophenol
there was "good agreement" with the statements given by examined persons.
     With respect to the possibility of information bias being present, this
could certainly lead to an enhancement of the risk ratios  as persons with STS's
-------
would more likely assume their disease was connected with exposure to herbi-
cides because of intense media publicity.  However, in a rebuttal  to the con-
clusions of the Australian Royal  Commission, Hardell reminded the  Commission
that debate about phenoxyacetic acid herbicides had abated with the banning of
2,4,5-T in Sweden in 1977, thus reducing media coverage (Axelson,  1986).
Hardell pointed out that the studies were undertaken during the period from
1978 to 1981.
     For the purpose of assessing the possible presence of observation bias in
his study, in 1981 Hardell repeated his analysis using colon cancer cases but
with the same controls as in the  first study.  Hardell  found no association
between exposure to phenoxyacetic acid herbicides and/or chlorophenols and the
risk of colon cancer.   He pointed out that this finding is inconsistent with the
occurrence of observational  bias  in the assessment of exposure (Hardell,  1981).
     The issue of lack of site specificity with respect to the occurrence of
STS is clearly not an  unexpected  phenomenon.  Carcinogens are not  always  site-
specific.   There are many examples of multiple site carcinogens (1,3-butadiene,
vinyl chloride,  and arsenic  to name a few).  Furthermore, the human body con-
sists of two main types of tissue, i.e.,  epithelial and connective.   Epithelial
tissue constitutes all  the duct glands, skin, the GI tract, neural  tube,  the
respiratory tract,  etc.  Malignant tumors originating in these tissues are called
carcinomas.  Connective tissue, on the other hand, consists of bone,  cartilage,
fat, muscle,  and subcutaneous tissue.  Malignant tumors originating in connec-
tive tissues  are called sarcomas.  With respect to appearance, a carcinoma is a
hard, adhesive mass with a well-defined,  advancing front.   Microscopically,  a
carcinoma of one organ is distinguishable from that of  another. In contrast,
sarcomas are soft,  jellylike, and fall apart easily. Microscopically,  they are
not clearly distinguishable  by site.  Corresponding to  the diverse  nature of
-------
connective tissue throughout the body,  sarcomas could be found anywhere  in  the
body.  They have poorly demarcated anatomic boundaries in contrast to carcino-
mas, and they contain large quantities  of intercellular materials.   As such,
they are subject to the same kind of insult as is the intercellular material  of
epithelial tissue.
     With respect to other significant  confounding factors in these studies,
there appears to be little substantiation of that criticism.   Harden certainly
controlled for age, sex, place of birth, and place of death.   He further analyzed
smoking habits and found them not to be different between the cases and controls.
Hardell does state, however, that other pesticides might constitute confounding
factors.  The Commission's statement regarding the presence of "other signifi-
cant confounding factors" is a quantum leap from Hardell's commentary about his
own data.  This appears to consist mostly of inuendo rather than substantive
criticism.
     The term STS is a convenient rubric under which all sarcomas arising out
of connective tissue have been classified.  Individual tumor types, i.e.,
rhabdomyosarcoma, fibrosarcoma, liposarcoma, etc., are subtypes that identify
a specialized connective tissue tumor.   In the Ninth Revision of the International
Classification of_ Diseases and Causes of_ Death, STS's are coded mainly to
category 171x; however-, under certain circumstances, they are coded to several
other diffuse categories as well.  If the STS is not classified to a specific
site,  it is given an ICD code of 171.9.  If the STS is coded to a site of the
body other than certain organ sites, it is still coded to the 171x category.
However certain STS's of specific organ sites are coded to.cancer of that site.
This could serve to reduce observed STS deaths that may be due to exposure to
polychlorinated dibenzo-p_-dioxin and spread them out over several site-specific
categories.  However, since expected deaths based on population death rates in
-------
category 171x would be subject to the same degree of difference as observed
deaths, one would not expect that the calculations of relative risks would be
affected although the ability to detect such elevated risks as significant
would be reduced.  Clearly, electron microscopy or use of immunohistological
techniques are needed to distinguish a sarcoma at one site from a sarcoma at
another site, but these techniques are rarely used.  This difficulty in diagnosis
is compounded by the fact that even collectively STS's are an exceedingly rare
form of cancer in the population.  Age-adjusted (1970) standard incidence rates
based on category 171x during the period from 1973 to 1977 were estimated by
the Surveillance, Epidemiology and End Results (SEER) group of the National
Cancer Institute to be 1.9 per 100,000 in the United States (Young et al.,
1982).  The corresponding age-adjusted mortality rate for ICD category 171x was
0.9 per 100,000 (Table 1).  STS rates generally follow the pattern of other
site-specific cancer rates by age, starting off low in younger age groups (less
than 2 per 100,000 in persons under 50) and increasing with age to a maximum of
12.2 per 100,000 in persons aged 85 and older.  Mortality rates remain under  1
per 100,000 up to age 50 but increase gradually to a high of 7.7 in persons age
85 and older.
     Additionally, the literature is replete with studies of carcinogens that
cause rare cancers.  These did not require confirmation before the scientific
community was convinced of their authenticity (e.g., DES and vaginal clear-cell
adenocarcinoma; vinyl chloride and angiosarcoma of the liver and glioblastoma
multiforme of the brain).
     Furthermore, although the Ninth Revision of the International Classification
of Diseases and Causes of Death has assigned a specific category (171x) to this
cause, it is not always certain that the correct diagnosis will be made.   This
was ably pointed out by Fingerhut et al. (1984) in a discussion of seven case
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   TABLE 1.  AVERAGE ANNUAL CRUDE, AGE-SPECIFIC, AND CUMULATIVE (age 0-74)
              INCIDENCE AND MORTALITY RATES FOR MALIGNANT CANCER
   INCLUDING in situ CASES BY PRIMARY SITE, ALL RACES, BOTH SEXES,  ALL AREAS
(EXCLUDING PUERTO RICO),
Age group
Crude
Adjusted
Cum (0-74)
<5
5-9
10-14
15-19
20-24
25-29
30-34
35-39
40-44
45-49
50-54
55-59
60-64
65-69
70-74
75-79
80-84
85+
Soft tissues
Incidence
1.9
1.9
0.2
1.2
0.4
0.5
0.8
0.9
1.1
1.3
1.5
1.5
1.9
2.7
3.5
4.2
4.3
7.0
8.0
9.3
12.2
(including heart)3
Mortality
1.0
0.9
0.1
1.1
0.5
0.2
0.3
0.3
0.3
0.4
0.6
0.7
0.7
1.4
1.8
2.0
2.9
3.5
4.1
4.7
7.7
1973-1977
Non-Hodgkin's
Incidence
8.9
9.0
0.8
0.6
0.9
1.0
1.0
1.6
2.3
2.7
3.6
6.1
8.9
13.6
18.9
27.3
34.8
45.8
51.3
53.2
51.7

lymphoma3
Mortality
4.9
4.8
0.4
0.2
0.4
0.4
0.6
0.5 ,
0.7
1.0
1.5
2.2
4.2
6.4
9.9
13.6
19.0
27.3
34.0
39.5
40.8
aper 100,000 population

SOURCE:  Young et al.,  1982.
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reports of STS.  Two of the seven were shown to be carcinomas.  This  difficulty
and the exceptionally rare nature of this tumor creates methodological  problems
in the assessment of risk.  As most trained epidemiologists  are aware,  the use
of the cohort technique to estimate risk of rare cancers is  not recommended.
Even the detection of not-so-rare cancer risks requires the  generation  of thou-
sands of person-years following the completion of a suitable latent period.
Hence, cohort analyses of rare cancers are inappropriate vehicles  for the
assessment of risk.  They lack power and are insensitive.
     The issue of latency is of importance here.  Several  studies  have  been
published that purport to show no association between exposure to  2,3,7,8-TCDO
and the risk of STS as well as to most site-specific cancers. Part of  the
explanation for this may be the short period of time that has elapsed between
onset of exposure and clinical manifestation of disease, in  this case,  cancer.
In most nonpositive epidemiologic studies, this period has been under 10 years.
This short period of time is probably inadequate to assess the risk of  most
site-specific cancers including STS.  Few cancers have been shown  to have  a
latent period under 10 years.  Hueper and Conway (1964) suggested  a latent
period for the development of sarcomas of between 15 and 30  years.  In  the
positive epidemiologic studies of Hardell and Sandstrom and Eriksson et al.,
the latent period has been found to range from 9 to 27 years after initial
exposure to the phenoxyacetic acid herbicides and/or chlorophenols with a
median time lapse of 20 years.  In the Lynge (1985) study, the lapse of time
from start of employment  (exposure) to diagnosis ranged from 14 to 26 years if
one excludes the one STS in that study with a lapsed period of time from initial
employment to diagnosis of only 5 years.  That is not likely to be due  to
exposure to 2,3,7,8-TCDD-contaminated chemicals.  If one assumes  that the
latent period based on these positive studies is between 14 and  27 years,  then

                                       8
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it is not likely that an excess risk of STS will  appear  until  at  least  the  15th
year after initial exposure.  This risk should increase  until  around  the  20th
year.  Probably 20 years should be considered the latent period for the develop-
ment of STS from exposure to 2,3,7,8-TCDD-contaminated chemicals  if one accepts
the hypothesis of a causal relationship which by  no means is  a certainty.
     It is thus remarkable that several cohort studies of individuals exposed
to 2,3,7,8-TCDD-contaminated herbicides and chlorophenols have produced any
STS's.  And, of course, the expectation of their  appearance is so small that
significance tests cannot be done.  The method of choice then is  the  case-
control technique.  However, the case-control technique  is also frought with
pitfalls, many of which are apparent in the evaluation of the epidemiology  data
on 2,3,7,8-TCDD.  With this technique, controls are matched with  actual cases
usually on the basis of known confounders (i.e.,  age,  race, sex)  or suspected
confounders (i.e., time of death,  length of employment,  and so on).   Uncontrolled
confounding, known or unknown, can play a significant  role in the determination
of a risk, as well as other factors such as latency or selection  of an  appropriate
index of exposure if exposure cannot be directly  measured.  The net result  can
lead the researcher to attribute a significant risk, if  found, to the exposure
being measured instead of to a confounder which may be the real culprit.
Second, if an index of exposure is used that does not  truly measure the actual
                                                                 f
exposure, misclassification will result.  This will force the risk estimate
toward the null.  Hence, if a true risk is present, it will not  be found.
Persons with no actual exposure could be classified as "exposed"  while  persons
with actual exposure could be classified as "unexposed."  Hence,  it is  always
better to choose a surrogate that is as close as  possible to the  target organ
                         /
dose.  For example, human tissue levels of 2,3,7,8-TCDD  would be  a far  better
surrogate of exposure than would information that a person was located  in or
-------
near a spot where 2,3,7,8-TCDD was  present.
     Biases can also increase or reduce the  risk  calculation depending on their
nature.  There is really no way to  control completely for some biases, i.e.,
recall bias and, to some extent, even information bias when dealing with ques-
tionnaire data no matter how hard the researcher  might try.  And  these will
always loom as potential problems in any case-control  study, no matter how well
conducted and designed it is.  On the other  hand, some biases could be elimina-
ted by a suitable selection of controls.  For example,  if  cancer  controls are
used, the researcher is obligated to eliminate those cancer controls where it
is suspected that the cancer might be associated with the  exposure under  inves-
tigation.  This is definitely within the power of the investigator.  Based on
the previous arguments  there  is no real basis upon which to conclude  that the
Hardell and Sandstrom (1979)  and Eriksson et al.  (1979,  1981)  studies  are not
good  scientific studies of the  risk of STS.
      Several other studies provided some support to the finding of an  excess
risk  of STS among individuals exposed to 2,4,6-trichlorophenol  (TCP)  and/or
2,4,5-T.  Zack  and Suskind  (1980), in a small cohort study,  noted that one
worker among 121 workers accidentially exposed to TCP (contaminated with  2,3,7,8-
TCDD) in 1949 developed STS's.  All  121 were chosen for this study because they
developed chloracne which indicates substantial  exposure to 2,3,7,8-TCDD.
No STS's would  have been expected.
      Lynge  (1985), in an incidence study of 3,390 males employed in two factories
manufacturing phenoxyacetic  acid herbicides, chiefly 2,4-D and MCPA,  found a
nonsignificant  excess risk  of STS  in male employees.  The author stated that
these results supported the Swedish  observation  of  an increased  risk of STS
following  exposure to phenoxyacetic  acid herbicides  "unlikely to be contaminated
with  2,3,7,8-TCDD."  However, after  a  10-year  latency,  the excess of STS was
                                        10
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significant (4 observed vs.  1.00 expected;  p  <  0.05)  in  male  employees of the
single factory where 2,4,5-T had been  produced  and  used  and where all five
STS's arose.  However, the author cautions  that because  of the  limited amount
of 2,4,5-T processed at that factory,  exposure  is  unlikely although  not  impossible.
     In a study of 2,189 manufacturing employees of a chemical  plant that
produced TCP, 2,4,5-T, and other chlorinated  phenols  contaminated with hexa-
chlorinated to octachlorinated dioxins, Cook et al. (1986) reported  one  STS
which was previously found by Fingerhut et  al.  (1984) in a case review,  through
pathology evaluation, not to be an STS.  However,  in a 1985 letter to the
Carcinogen Assessment Group of the U.S. Environmental Protection Agency,
Cook reported that another member of the cohort was found to be suffering  from
a confirmed STS.  This case was not reported as an STS because  the person  died
after the end of the follow-up period.  Cook et al. did not estimate how many
deaths should be expected.  Latency was not considered although it appears  that
a sufficient follow-up had been achieved in which latency could have been
evaluated for the risk of cancer at other sites.  The authors believe that
their data, provide little evidence of a cancer risk from exposure to 2,3,7,8-
TCDD especially STS.
     In a later update of this same cohort by Ott et al. (1987), the authors
employed a mathematical analytical technique known as a "serially additive
expected dose model" designed by Smith et al. (1980).  This statistical  device
has several  limitations just as does the commonly accepted method of simply
analyzing for latent effects by calculating  observed and expected mortality
after a lapse of a sufficient number of years from initial exposure.  Despite
the subjective designation  of jobs according to an "intensity of exposure"
scale from 0 to 4, which is not based on actual exposure and is subject to
misclassification bias, it  still takes 15 to 20 years for the development and
                                       11
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detection of a malignancy!  If few persons in the  cohort have achieved that
latent period, no excess cancer risk will  be detected no matter what the
"intensity" of exposure.  This method also assumes that one year of exposure at
a high intensity score is equivalent to 10 years of exposure  at the next lower
score in its ability to produce cancer. This is not a valid  assumption because
such an equivalency fails to consider that a brief excursion  at a  very intense
exposure could affect the metabolism differently from that of a long continuous
(or discontinuous exposure).  This device, although limited  in some respects,
allows for the assessment of latent effects for certain sites.  Only total cancer,
stomach cancer and lymphomas were analyzed for latent effects in  this manner
due to the "paucity of deaths through 1987."  Except for total malignant neo-
plasms, only six stomach cancer deaths and six lymphomas were separately analyzed.
The authors saw no trend of increasing risk of cancer at the  few  sites examined.
However, it must be kept in mind that with a long latent period  required for
the manifestation of cancer, the power to  detect a significant risk of cancer
at these two sites must of necessity be small, even for  this  technique, although
the authors did not estimate it.  Of course, no other cancer  sites were analyzed
for latent effects by this method.  Of interest is the persistently high signi-
ficant risk of "other and unspecified" cancer (12  observed vs. 4.6 expected,
C.I. 135-456) for which the authors could offer no explanation.  This cohort  or
some portion of it wil be included in the  National Institute  for  Occupational
Safety and Health (NIOSH) registry of exposed workers that will  be analyzed by
NIOSH in the near future.  One question remains, however,  and that is, if
2,3,7,8-TCDD tissue levels in a substantial portion of these  workers are similar
to 2,3,7,8-TCDD tissue levels in the background referent population and always
have been (thus indicating little evidence of differential exposure), an elevated
risk of cancer would not be expected to occur.  It is possible that due to the
                                       12
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7-year half-life of 2,3,7S8-TCDD  (Pirkle  et al.,  1987), tissue levels that were
elevated in the past from early exposure  to 2,3,7,8-TCDD  have returned to
background levels,  but of course,  there is no  evidence that tissue levels of
2,3,7,8-TCDD in the past were ever elevated  (not  always the same  as they are
today), and that possibly no elevation ever occurred even  during  times when
potential exposure was hypothesized.
     Hence, because of this possibility,  and the  knowledge that the presence of
chloracne at some time in the past more than likely  is proof  that a substantial
exposure to 2,3,7,8-TCDD occurred, it seems likely that a  cohort  consisting en-
tirely of chloracne persons having known  contact  with 2,3,7,8-TCDD in the past
would provide the most ideal cohort for assessing the long-term health effects of
exposure to TCDD!  This is not the same as  saying that the presence of chloracne
is.necessary in order to have exposure.   The major problem with current efforts
to determine tissue levels of 2,3,7,8-TCDD  is  that  they  don't reveal what the
tissue levels were in the past.
     Smith et al. (1984) conducted a case-control study  of STS  in New Zealand
and found a significantly high risk of STS  among  railroad workers.  Specific
exposure to phenoxyacetic acid could not  be  identified.   Tannery  and meat
workers were also found to be at a significantly  high  risk of STS.   In those
jobs among meat workers where pelts are treated with chemicals  such as TCP
(2,4,6-trichlorophenol) that contain 2,3,7,8-TCDD,  the  risk  ratio increases  but
is marginally significant because of small  numbers.  Although Smith et al. found
excess nonsignificant  risks in applicators and users who were exposed  to  only
phenoxyacetic acid herbicides and/or chlorophenols,  their personal exposure
information is not substantiated, similar to the  Swedish studies.  In  a later
update of this study (Smith and Pearce,  1986), which included more cases  and
controls that had been diagnosed, the excess was  reduced to  the null.  The
                                       13
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 authors have been  criticized  by Axel son  (1986) for failing to exclude from the
 referent group cancers that may be  related to the exposure in question (i.e.,
 lymphomas),  thus potentially  underestimating the true risk.  Smith et al.
 concluded that at  the levels  of exposure experienced by individuals in contact
 with  ground  spraying,  it  is unlikely that 2,3,7,8-TCDD could cause STS.
      Other studies that are consistent or are not inconsistent with the finding
 of  an elevated risk  of STS  among persons potentially exposed to polychlorinated
 dibenzo-£-dioxin-contaminated herbicides and/or chlorophenols are Balarajan and
 Acheson (1984), Cantor (1982), Milham (1982), Kogan and Clapp (1985), Michigan
 Department of  Public Health (1983a, b), Fett et al. (1984), and most recently,
 Puntoni et al.  (1986), Merlo  and Putoni  (1986), Woods et al. (1987), and Kang
 et  al.   (1987).  However, in  many but not all of these studies, definite evidence
 of  exposure  to 2,3,7,8-TCDD is not established.  Furthermore, the risk estimates
 are based  upon  generally  small numbers as one might expect.
      In a  case-control  study  based on data provided by the National  Cancer
 Register of  England and Wales, Balarajan and Acheson (1984) found a  significant
 risk  of STS  among  farmers,  farm managers, and market gardeners, but not among
 agricultural  workers,  gardeners, and groundsmen.  The authors hypothesize exposure
 to  2,3,7,8-TCDD-containing  herbicides in high risk occupational groups as a
 possible cause, although actual  evidence of exposure is not provided.
     Cantor  (1982), in  a case-control study,  noted a significant risk of  reti-
 culum cell sarcoma among farmers under age 65 in Wisconsin in counties that
were high  in suronary measures of General  Agricultural  Activity based on county-
wide measures of farm-related exposures.   However,  occupations had been clas-
 sified  based on what was recorded on the death certificates.   Often,  the  infor-
mation is absent,  however, or if an  occupation is mentioned,  it is frequently
the last job  held or the job of  longest duration or simply the .word  "retired"
                                       14
-------
will appear.  Such information must be considered unreliable  at  best and  could
lead to misclassification.  Furthermore,  the use of 2,3,7,8-TCDD-containing
chemicals is only suggested by the authors.
     Mil ham (1982), in a proportionate mortality rate (PMR) study,  noted  a
significantly high risk of STS among farmers in Washington State.   The author
states  that exposure would most probably have been to phenoxyacetic acid  herbi-
cides  (chiefly 2,4-D) and chlorophenols.   There are two major problems with
this study:   (1)  no actual evidence of exposure is provided,  and (2)  the  source
of  death  information is not given.  Additionally, PVR studies are inherently
inferior  to cohort and case-control designs.
     Kogan and Clapp  (1985),  in a second PMR study, found a significantly high
rate of connective tissue cancer  (STS) among Vietnam veterans in Massachusetts
as  opposed to non-Vietnam veterans  in Massachusetts.  Again, service in Vietnam
appears to  be the surrogate of exposure and as  such may bear no relationship to
actual  exposure  to 2,3,7,8-TCDD-containing chemicals.  It  is entirely possible
 that many Vietnam veterans  in Massachusetts were never exposed  to  large  quanti-
 ties of 2,3,7,8-TCDD-containing chemicals, while  there is  no reason to think that
 some non-Vietnam Massachusetts veterans may have been exposed to 2,3,7,8-TCDD-
 containing chemicals in  other settings such as  occupational.
     The  Michigan Department of Public Health  (1983a, b)  found  significantly
 high death rates among females of Midland County, Michigan,  compared  to  the
 national  average for the period 1960 through 1978.  In  contrast, STS  death
 rates  for males of Midland County were not elevated during the  same period.
 The Dow Chemical Company produced phenoxyacetic acid herbicides and/or chloro-
 phenols  in this  county.   Since this is an ecological study,  it  will of necessity
 include  in both the numerator and denominator individuals who may  or  may not
 have  received exposure to 2,3,7,8-TCDD or polychlorinated dibenzo-p_-dioxin-
                                        15
-------
contaminated chemicals.  Cook and Cartmill (1984), in a discussion  of the  soft
tissue sarcoma issue, maintained that the Michigan Department  of Public  Health
found no commonalities among the women that would suggest any  single agent
including exposure to 2,3,7,8-TCDD-contaminated chemicals as the cause except
that several of the women moved to Midland County with pre-existing diseases.
Whether the diseases they brought with them had anything to do with STS  was  not
established.  The authors also suggested that several women may have been
incorrectly classified to ICD category 171x since the site of  the cancer was
not specifically mentioned.  However, there is no reason to suspect that the
Michigan Department of Public Health coded their death certificates any  dif-
ferently than did the National Center for Health Statistics, the U.S. agency
responsible for generating U.S. and state vital statistics.  These death rates
provide the basis for calculating the expected deaths in that  study.
     A study of Vietnam veterans in Australia (Fett et al., 1984) essentially
reported nonpositive findings, but the veterans also exhibited an excess of STS,
albeit small (i.e., 2 observed vs. 0.64 expected).  In the comparison group of
non-Vietnam veterans no STS's were observed versus 0.84 expected.  This  study
again used "service in Vietnam" as a surrogate for exposure to Agent Orange.
No actual proof of exposure is provided, and one would expect many if not  most
of the Vietnam veterans to have had little exposure to Agent Orange.
     A recent study (Puntoni et al., 1986; Merlo and Puntoni,  1986) reports a
high incidence of STS in residents of the region around Seveso, Italy, where an
industrial explosion at a factory that produced 2,3,7,8-TCDD-contaminated
2,4,5-T occurred on July 10, 1976.  Few specifics are available regarding  this
study, but it is apparent that the rate of STS's was higher in residents of the
Seveso region as contrasted with those in the Varese region (2 per 100,000)  and
tended to increase with time (i.e., 4.35 in 1981) when so-called "polluted"
                                       16
-------
areas are lumped with "non-polluted"  areas.   The authors  suggested that these
two areas be combined since chloracne rates  calculated for each area separately
were inconsistent with the defined soil  levels  of 2S3,7,8-TCDD in each area,
the surrogate measure used for defining  a "polluted"  area as contrasted with a
"non-polluted" area.  Misclassification  of exposure was  suggested by the  authors
as the explanation for the lack of a  positive correlation with the incidence of
STS in so-called "polluted" areas.  Hardell  and Eriksson (1986) suggested that
exposure to the chemicals produced by this factory prior to the accident
accounted for the high rates befores  during, and after the  accident  in the
Seveso region since latent factors argue against the accident being the cause
of the excess.
     The remaining nonpositive studies are even less remarkable than those
already mentioned.  Thiess et al. (1982) examined a cohort  of 74  employees of
the Badische Anil in und Soda Fabrik (BASF) who  were accidentally  exposed  to
trichlorophenol in 1953, 66 of whom suffered chloracne;  after a lengthy follow-
up, no STS's appeared.  However, the study has  little power to detect  a sig-
nificant risk of cancer.
     Axel son et al. (1980) studied 348 railroad workers  who sprayed  2,4-D and
2,4,5-T herbicides from 1957 to 1972 and found  no STS.   Again, this  study has
limited power to detect a risk of STS as well as little evidence  to  substantiate
exposure.
     Riihimaki et al. (1982) studied 1,926 Finnish applicators of 2,4-D and
2,4,5-T from 1955 to 1971 and found no STS's (0 observed vs.  0.1  expected).
However, the authors suggested major difficulties with their paper as  follows
". . . an observation period under 10 years, the presence of  selection bias,
and low or little exposure."  They conclude that it "cannot be  regarded as a
negative study."  This study also suffers from  survivorship bias.
                                       17
-------
     Wiklund and Holm (1986) recently studied a massive  cohort  of 354,620
Swedish men who were recorded as  having an agriculture or forestry job according
to the census of 1960 versus 1,725,845 Swedish men in all other industries.
The primary exposure in those jobs was postulated  to be  MCPA; 2,4-D and 2,4,5-T
were also used to a lesser extent.  The authors found that  the  relative risk of
STS was only 0.9.  This study has several  problems:   (1) a  lack of individualized
exposure data; (2) only 15% of Swedish agricultural  and  forestry workers were
estimated to be exposed to phenoxyacetic acids and 2% to chlorophenols; (3)
Swedish agricultural workers are known to have a decreased  cancer  risk and  use
health services less frequently;  (4) classifying workers according to a 1-year
employment status presents the possibility of misclassification; and  (5) the
crude rate of STS in agricultural and forestry workers based  on data  in the
study is 5.45 per 100,000 person-years, and in the remaining  workers  it is  5.00
per 100,000 person-years.  Both rates are high compared  to  rates from other
nations (1 to 3 per 100,000 person-years).
     The authors of the previous study, Wiklund et al.  (1987),  recently evaluated
the relative risks of Hodgkin's disease and non-Hodgkin's lymphoma in a cohort
of 20,245 Swedish pesticide applicators who were licensed anytime  between  1965
and 1976.  The authors estimate that 72% had had an opportunity for contact
with phenoxyacetic acid herbicides (chiefly MCPA, mecoprop, and dichloprop, and
to a lesser extent, 2,4,5-T and 2,4-D).  This estimate was  based on responses
to a questionnaire that had been mailed to a random sample  of 273  persons  in
the cohort.  No actual measurements of exposure were taken.  Overall, based on
a rather short follow-up averaging 12.2 years, no significant excess  risk  of
either disease was found.  However, the authors noted a  trend of increasing
risks for both diseases with lapsed time since licensing.   Further follow-up  of
this cohort is needed since presumably the latent period for lymphoma is  in
                                       18
-------
excess of the observation period of this study.
     Another nonpositive study (Lathrop et al., 1984a,  b;  Wolfe  et  a!.,  1984,
1985) of a group of young Air Force officers and skilled and unskilled enlisted
men (called Ranch Handers) who presumably were  involved with the aerial  spraying
of herbicides containing 2,3,7,8-TCDD revealed  no excessive cancer  risk.  How-
ever, only 6 of 1,256 Ranch Handers actually developed  cancer in the short
period of time they were followed.  Although this study offers little evidence
to substantiate exposure to Agent Orange by itself, recently Pirkle et al.
(1987) reported that some 75 Ranch Handers (6% of the total) had reported a
range of exposure to 2,3,7,8-TCDD (from 16 to 423 ppt)  based on  adipose tissue
samples collected by the Centers for Disease Control (CDC) and the  Air Force.
Furthermore, the study suffers from a lack of power and short latency.
     Another nonpositive study by Greenwald et al. (1984) of 281 STS cases from
the New York State Cancer Registry matched with 281 referents taken from New
York driver's license files and 130 from New York death certificate files
exhibited only a nonsignificant excess cancer risk for workers in chemical
manufacturing and highway construction.  Based on questionnaire responses, the
risk from exposure to 2,3,7,8-TCDD-contaminated Agent Orange was only 0.7 while
that of herbicide and/or pesticide use was 1.0.  The risk for farming was 0.79.
Two major problems are evident in this study.  First, "service in Vietnam" was
used as a surrogate for exposure to Agent Orange.  This potentially can lead
to a problem similar to that of other studies where exposure to Agent Orange
could  not be substantiated, and secondly, those who did not serve in Vietnam
could  have  received exposure to 2,3,7,8-TCDD through other means.  Furthermore,
the  average time lapse from Vietnam service to a diagnosis  of STS in this study
was  only 11 years, a short latent period.  Many additional years of observation
must be added before the latency question can be put to rest.
                                       19                         /
-------
     Kang et al. (1986) group-matched 234 Vietnam-era veterans who were STS
patients and had served in the U.S. military between 1964 and 1975 with 13,496
patients sytematically sampled from the same Vietnam-era patient population
without a diagnosis of STS.  Service in Vietnam (a surrogate for exposure)  did
not appear to be associated with STS.  However, the authors reported that more
than likely most patients (and controls) had not achieved a follow-up time
sufficient for latent effects to become manifest.  Furthermore, actual exposure
as evidenced by in vivo measurements of 2,3,7,8-TCDD were not available or, as
the authors pointed out, Vietnam veterans may have been exposed to such small
amounts that standard epidemiologic designs could not detect an excess risk.
     In a second case-control study, Kang et al. (1987) selected 217 STS
patients from the Armed Forces Institute of Pathology and matched them to 599
matched controls for service in Vietnam, exposure to chemicals, medical history,
and life style.  Military service was verified by the patients'  military per-
sonnel records.  Vietnam veterans, in general, did not exhibit increased risks
of STS.  However, the authors found that the risk of STS increased in veterans
with combat experience versus those without (OR = 2.57, nonsignificant) and
increased even more for those same combat veterans who were assigned to the
region where most spraying of Agent Orange took place (OR = 8.64, nonsignificant).
The authors noted that their study had "very low power" to detect enhanced
risks in subgroups of veterans who had greater opportunities for exposure to
Agent Orange.  The authors concluded that a possiblity exists that certain
subgroups of Vietnam veterans may be subject to "modestly increased risks of
STS" from exposure to Agent Orange, but it can neither be confirmed nor ruled
out.
     Karon et al. (1987, unpublished), at a recent meeting on dioxin, reported
only one army veteran with a current serum 2,3,7,8-TCDD level in excess of
                                       20                        -~
-------
20 ppt out of some 775 veterans who participated in  the study.   About  675  of
these served as ground combat troops in Vietnam and  included some of the
troops who presumably had had a high likelihood of exposure to  Agent Orange.
These levels today are not different from those found in background populations
of industrialized nations.  At first glance this seems to argue that veterans
in Vietnam may not have been heavily exposed to Agent Orange,  and hence findings
based on the surrogate experience in Vietnam may not be relevant.  Such may
not be the case since actual time served in Vietnam  occurred a considerable  num-
ber of years ago.  Elevated serum 2,3,7,8-TCDD levels brought on by massive
exposure to Agent Orange at that time may have dissipated and returned to
background levels during the period following removal of troops from Vietnam.
     The Hoar et al. (1986) study is not contradictory to the hypothesis that
2,3,7,8-TCDD is the contaminant responsible for the  development of STS. The
authors found that 2,4-D, which does not contain 2,3,7,8-TCDD,  was identified
by most members of the cohort as being the herbicide to which they were
exposed.  Again, the results of this study are based on questionnaire  data and
may suffer from recall bias.  This study supports only the hypothesis  that the
herbicide 2,4-D appears not to be a credible candidate for the carcinogenic
agent responsible for STS.

                             NON-HODGKIN'S LYMPHOMA

     Hardell et al.  (1981), in a case-control study  of lymphomas in Swedish
workers, found a statistically significant risk of lymphoma in  agricultural,
forestry, and woodworking employees from exposure to phenoxyacetic acid herbi-
cides and/or chlorophenols.  Risk ratios ranged from 4.3 to 6.0 for both classes
of compounds together as well as separately.  However, exposure to organic
                                       21
-------
solvents such as benzene, trichloroethylene,  and styrene was  also found to be a
risk factor.  Unfortunately, the data were based on answers derived  from adminis-
tered questionnaires, and the exact extent of exposure  to  2,3,7,8-TCDD is
difficult to determine.  As such,  this study  is subject to some recall bias and
confounding, but the risk ratios are sufficiently high  that problems with the
study probably could not account for them entirely.   This  is  another one of the
Swedish studies that had been severely criticized for alleged biases, confound-
ing, and distortions.  The discussion following the Harden and Sandstrom
(1979) and Eriksson et al. (1979,  1981) studies regarding  those criticisms
pertains to this study as well.
     Hoar et al. (1986), in a similar study,  found significantly high rates of
non-Hodgkin's lymphoma (NHL) in farmers in Kansas who use  herbicides, particu-
larly 2,4-D and triazines.  Few respondents could remember exposure  to 2,4,5-T
Which contains 2,3,7,8-TCDD.  2,4-D is not believed to  contain 2,3,7,8-TCDD but
does contain other polychlorinated dibenzo-p_-dioxin impurities. The risk was
found to increase with increasing frequency and duration of herbicide usage.
Although "herbicide usage" could mean any of  the herbicides identified by Hoar,
she wrote that this is "essentially synonymous" with use of 2,4-D.   The next
most used herbicides (i.e., triazines and uracils) are  nonsignificant when
exposures to phenoxyacetic acids are controlled for. However, this  study has
problems similar to the Harden et al. (1981) study in  that there is a lack
of substantiation of exposure, and the information is based on questionnaire
responses that are subject to some recall bias.  Moreover, there is  a statisti-
cally significant risk associated with the use of other herbicides as well,
i.e., triazines, amides, and trifluralin.
     A population-based case-control study of the relationship of occupational
exposure to the risk of STS and NHL was recently completed by Woods  et al.
                                       22
-------
(1987).   A total  of 128 STS cases  and 576  NHL  cases  diagnosed  between  1981 and
1984 were matched with 694 randomly selected controls without  cancer.  The
authors  reported no overall increased risk of  either disease from past
occupational exposure to phenoxy herbicides or chlorophenols based on  personal
interviews.  However, significant  elevated risks  of  NHL  were observed  in  certain
subgroups thought to have somewhat heavier exposure  to herbicides,  i.e.,  farmers
(OR = 1.33, CI 1.03-1.7), forestry herbicide applicators (OR = 4.8,  CI 1.2-19.4),
and persons potentially exposed to phenoxy herbicides for 15 or more years
during the period 15 years prior to cancer diagnosis (OR = 1.71, CI  1.06-1.7).
Of interest in this study is the finding that presence of chloracne is associated
with a high (borderline significance, p <  0.075)  risk of STS and less  so  to  a
risk of NHL.  In so far as the information on exposure,  based  on questionnaire
data, is somewhat suspect because  actual confirmation was not  obtained (i.e.,
adipose tissue specimens with confirmed presence of  2,3,7,8-TCDD),  the presence
of chloracne in conjunction with known contact with  chemicals  containing  2,3,7,8-
TCDD could have meant a massive exposure to 2,3,7,8-TCDD above and beyond
background levels.  The chloracne  was not  clinically confirmed.  Because  the
use of questionnaires to elicit past-history of exposure to specific phenoxy
herbicides and/or chlorophenols has a potential for  misclassification  among
cases and controls  (because of memory problems and lack of confirmation of
exposure), this study seems to have reported paradoxically conflicting results.
     Buesching and Wollstadt (1984), in a case-control study,  found that white
male farmers of Winnebago County,  Illinois, had a statistically significant
risk of 2.65  for NHL as  contrasted with all white males of that county.   Although
the author presents no evidence of exposure to any particular  herbicide or
chlorophenol, he suggested that phenoxyacetic acid herbicides  were probably
used.  Of major concern  in this study is the use of occupation as coded on the
                                                                 /•"
                                       23
-------
death certificates to classify the deceased by exposure.   As mentioned pre-
viously, use of such information may lead to misclassification.
     Cantor (1982), in a case-control study of NHL, found a significantly  high
risk of NHL among farmers under age 65 in counties of Wisconsin  known to be
high in insecticide and herbicide use, and having small  grain acreage, wheat
acreage, and dairy sales.  The risk of reticulurn cell sarcoma is higher at 3.2
among younger farmers in these same counties.  For small  grain acreage and
acres treated with insecticides, the risk is 6.6, and for wheat  acreage, it  is
4.4; all are significant.  Unfortunately, information on  exposure was derived
based on occupations listed on death certificates similar to that of Buesching
and Wollstadt (1984) and as such cannot be considered reliable.   The potential
for misclassification of exposure among the cases and controls is great.
Additionally,  no particular herbicide, insecticide, or pesticide was mentioned.
They were only surmised to be in use by the author.
     Burmeister et al. (1983), in another case-control  study of  NHL  in Iowa
farmers, found a statistically significantly high risk of NHL associated with
farming.  It was found to be elevated in those born before 1901  and  was associ-
ated with jobs involving egg-laying chickens, solid milk  products, hog produc-
tion, and herbicide use.  This study, which also depended on occupation as
recorded on death certificates for exposure, may be very  unreliable, and it  is
likely that misclassification could have resulted.  The authors  suggested  that
herbicides and/or insecticides, such as those mentioned in the Hardell and
Sandstrom (1979) study, may be carcinogenic.
     Milham (1982), in a proportionate mortality ratio study of  deceased workers
in the agriculture, forestry, and wood products industries of Washington State,
noted a significant excess of Hodgkin's disease and a high but nonsignificant
risk of NHL in paper and pulp mill  workers.   Again, there is no  evidence of
                                       24
-------
actual exposure to phenoxyacetic acid herbicides  and/or  chlorophenols  cited by
the authors.   Furthermore,  the proportionate mortality  design is inherently
inferior to the cohort or case-control  study design because the former forces
interdependence of the risk estimates.
     Cook et al. (1986), in a cohort study of 2,189 manufacturing employees of
a chemical plant that produced TCP, 2,4,5-T, and other chlorinated phenols
found a nonsignificant increased risk of NHL (5 observed deaths versus 2.1
exposed).  However, the authors reported that these results do not support  a
causal association between chronic disease and exposure to the chemicals in
question.  The problems with this study involve lack of evidence of exposure
to most members of the cohort  (only 15% had chloracne), and no effort was made
to assess latent effects.
      Ott et al. (1987) updated the Cook et al. (1986) study by using a
statistical technique known  as a "serially additive expected dose model" that
assesses latent effects.  Based on six lymphomas,  no latent trend was found.
The  problems with  this  design were discussed previously.
      Other studies, such as  those of Lynge  (1985), lack and Suskind (1980),
Thiess  et  al.  (1982), Axel son  et al.  (1980), Riihimaki  et  al.  (1982), Lathrop
et al.  (1984a,  b), Wolfe et  al.  (1984), Fett et al.  (1984), and Kogan and Clapp
 (1985),  did  not find  an excess risk  of NHL  in  individuals  who  were potentially
exposed to 2,3,7,8-TCDD-contaminated chemicals.  All of these  studies have
methodological  flaws  that  make it  unlikely  that  they could detect a significant
 risk if one were  present.  The Lynge (1985)  study  noted only a slight excess of
 lymphomas in  males after a lapse of  10 years  from  initial  exposure; however,
 sensitivity was somewhat reduced.  The  Zack  and Suskind (1980), Thiess  et al.
 (1982),  and  Axelson et al.  (1980)  studies were exceptionally small studies and
 hence lacked  sensitivity because of  their size despite  a  lengthy  follow-up.
                                        25                            /,''. •
-------
The Riihimaki et al. (1982) study, although seemingly large,  had many methodo-
logical deficiencies as described in the section under soft tissue sarcomas
that precluded detection of a risk of NHL.  The study by Lathrop et al.  (1984a,
b) and Wolfe et al. (1984) of Ranch Handers also lacked sufficient statistical
power to detect a risk as being significant.  Furthermore,  there was a lack of
evidence of substantiation of differential exposure to Agent Orange in this
study, a short latent period in which to expect the development of NHL,  and a
lack of sensitivity for the detection of NHL.  The Australian Veterans Health
Studies (Fett et al., 1984) are much like the Ranch Handers studies in that they
lack sensitivity, provide little evidence of differential exposure, and  have a
short latent period.  Kogan and Clapp (1985) did not evaluate NHL  in their
study of Massachusetts veterans.  This study also had several limitations which
were discussed in the section on soft tissue sarcomas.

                                 STOMACH CANCER
     Two small cohort studies by Thiess et al. (1982) and Axel son et al.  (1980)
 reported statistically significant excesses of stomach cancer (albeit three
 cases in each study), while a third smaller study (121 persons) by Zack  and
 Suskind  (1980) reported no excess risk of stomach cancer.  This exceptionally
 small cohort lacks sufficient power.  Only 0.5 stomach cancer deaths would have
 been expected to  have occurred by the end of the cut-off date in that study.
 Lynge (1985) reported a nonsignificant excess risk of stomach cancer in  chemical
 workers  in Denmark.  Cook et al. (1986) noted a slightly elevated risk of stomach
 cancer in his cohort which was also nonsignificant.  Cook et al., however, did
 not analyze their data for latent effects.  The Riihimaki et al. (1983)  cohort
 study is another  nonpositive study with large deficits of mortality.  This
                                       26
-------
study suffers from major problems not  the least  of which  is a survivorship
phenomenon.  The members of this  cohort  had to live  long  enough to be included
in the cohort.  In the Wolfe et al.  (1984) and Lathrop  et al. (1984a, b) study
of Ranch Handers, no excesses of  any cancer of any kind were found.  Again,
this mortality study of mostly young men produced only  six cancer deaths to
date, and has several  problems as mentioned previously.   Burmeister et  al.
(1983), in a case-control study,  noted a significant risk of stomach cancer  in
individuals who had jobs in "milk products, cattle production,  and/or jobs
involving high yields of corn per acre."  The authors suggested as a cause
possible exposure to herbicides and/or insecticides  such  as those  reported  in
the Hardell and Sandstrom (1979)  study.   However, occupation as recorded on
death certificates was the surrogate used by the authors  for exposure,  and  as
such is a very unreliable vehicle for  the determination of exposure.

                               ALL OTHER CANCERS

     Other cancer sites have been found to be significantly elevated among
persons or groups potentially exposed to 2,3,7,8-TCDD-contaminated chemicals.
Buesching and Wollstadt  (1984) and Burmeister et al. (1983) reported significant
elevated  risks of prostate cancer in farmers in Winnebago County,  Illinois, and
agricultural workers in  Iowa, respectively.  Both used death  certificate desig-
nation  of occupation as  their surrogate  for exposure.  As was  pointed out
earlier,  occupation as  recorded on death certificates is  very  unreliable.   No
substantiation of exposure to any herbicides or pesticides was done.
     Milham  (1982) found a significant excess of Hodgkin's disease in persons
employed  in  the  agriculture,  forestry, and wood products  industry of Washington
State.  His  source of data was not  revealed.  However, there is again no evidence
                                        27
-------
 actual  exposure to 2,3,7,8-TCDD-containing phenoxyacetic acid herbicides and/or
 chlorophenols  in this proportionate mortality study.
      Kogan and Clapp  (1985),  in his proportionate mortality study of Vietnam
 veterans  versus non-Vietnam veterans in Massachusetts, found a significant
 excess  of kidney cancer.  Again, no evidence of actual exposure to Agent Orange
 was  given.   This study and its methodological flaws were discussed previously.
 In short, service  in  Vietnam  is the surrogate for exposure.
      Finally,  Cook  et al. (1986) found a statistically significant excess of
 other and unspecified cancers in his study of employees of a chemical plant
 that produced  TCP,  2,4,5-T, and other chlorinated phenols.  Again, the flaws
 in this design were discussed previously.

                          .  SUMMARY AND CONCLUSIONS

      The  evidence that persons who are exposed to phenoxyacetic acid herbicides
 and/or chlorophenols are subject to an elevated risk of STS's is based primarily
 on two independent  case-control  studies by Harden and Sandstrom (1979)  and
 Eriksson  et  al.  (1979, 1981).  The problems with these studies, as outlined in
 the  section  on  STS, are not sufficient to explain the highly significant risks
 of STS found in workers exposed to these chemicals.  Power considerations alone
 indicate  that  as many as 1/3 of the positive responses in the cases could be
 reversed without a major loss of power in both studies.  Additionally, Eriksson's
 data  further suggest that the risk is greater from exposure to 2,3,7,8-TCDD-
 contaminated herbicides such as  2,4,5-T rather than from those herbicides free
 of 2,3,7,8-TCDD contamination.  Several additional studies tended to support or
were consistent with the Hardell  and Sandstrom (1979)  and Eriksson et al.
 (1979, 1981) findings of an excess risk of STS in certain subgroups of the study

                                       28                       "-
-------
groups thought to be more highly exposed to 2,3,7,8-TCDD-contaminated chemicals
(Zack and Suskind, 1980; Lynge,  1985;  Balarajan and Acheson,  1984; Cantor,
1982; Mil ham, 1982; Kogan and Clapp,  1983;  the Michigan  Department of Public
Health, 1983a, b; Puntoni et al., 1986;  Merlo and Puntoni,  1986; Wood et  al.,
1987; and Kang et al., 1987). However,  except for the Zack and Suskind study,
where proof of substantial exposure to 2,3,7,8-TCDD-contaminated  trichlorophenol
is evident by the occurrence of  chloracne in all  subjects,  differential exposure
to 2,3,7,8-TCDD, sufficient to produce a detectable change in risk  estimates,
can only be assumed to have occurred where proof of exposure is available.   In
addition, authors of several studies have maintained that their own  results  do
not support the results of the Swedish studies.  However, many of these studies
exhibit small nonsignificant risks of STS when it could be measured.  For
example, the Cook et al.  (1986)  study produced one STS where none were expected,
while the Smith et al.  (1984) study produced a significant risk of  STS  in
workers involved in treating pelts with chlorophenols.   Even the  Fett et  al.
(1984) study of Australian Vietnam veterans produced two STS's where only 0.67
were expected.  On the other hand, in studies where no  risk of STS  was  found,
such as Thiess et al.  (1982) and Axelson et al.  (1980), the cohorts are  exception-
ally small, and one would not expect to see any of the  exceedingly  rare STS's
even though many years  have  passed.  Even  the defective study of  Riihimaki et
al.  Il982) anticipated only  0.1 STS.  Wolfe and Lathrop's Ranch Handers exhibited
little cancer mortality,  and never achieved a latent period long  enough to
detect any STS's.  Wiklund and Holm (1986), on the other hand, in their massive
study of STS  in Sweden,  although nonpositive, provided enough data to calculate
crude STS rates that appear  unusually high compared to that of other countries.
Woods et al.  (1987)  found a  borderline  significant risk of chloracne although
clinically unconfirmed with  risk of STS.   Kang et al. (1987) found a high risk
                                        29
-------
of STS in Vietnam veterans most likely to have  been  exposed  to Agent Orange,
although power was very low.  Greenwald et al.  (1984)  used  "service in Vietnam"
as a surrogate for exposure.  Only five of his  nine  "exposed" Vietnam service
cases actually thought they were exposed to Agent Orange, while  four "non-Vietnam
service" controls claimed exposure to phenoxyacetic  acid herbicides in occupational
settings.  The Hoar et al. (1986) study seems to imply only  that 2S4-D and
triazines could be ruled out as a cause of STS, since  2,3,7,8-TCDD-contaminated
2,4,5-T was believed not be present.   Hence, this study does not really  contradict
the Swedish studies because 2,4,5-T was not considered.  Most of these studies
suffer from one or more methodological flaws that preclude  any  conclusion that
they are nonsupportive of a finding of a risk of STS.
     The net result is that basically the Swedish studies of Hardell and
Sandstrom and Eriksson et al. provide the best evidence of  a causal relationship
from exposure to the phenoxyacetic acid herbicides and/or chlorophenols.  There
is some additional support from some small cohort studies and a few poorly  and
not so poorly designed case-control and proportionate  mortality  studies.  However,
these two Swedish case-control studies, contrary to  the criticism leveled at
them, appear to have been constructed and executed along classic lines as
outlined in any good epidemiology course on methodology.  The authors  appear to
have been concerned about possible bias, confounders,  and other problems in
their analyses as well as their critics.  Harden and  Sandstrom re-examined their
data by using colon cancer cases and matching them with the same controls in
order to assess the effect of certain biases with which they were concerned.
Under any other circumstances these studies would be considered reasonably  good
studies minus certain analyses relating to dose response that probably were not
envisioned at the time by the authors.  Largely on the basis of these Swedish
studies, the International Agency for Research on Cancer decided to  reclassify
                                       30
-------
the phenoxyacetic acid herbicides and/or chlorophenols in 1986 as probable
human carcinogens based on 1 invited evidence of their carcinogem'city  in  humans.
     Although the epidemiologic data appear to provide limited evidence  that
exposure to phenoxyacetic acid herbicides and/or chlorophenols is causally
related to the risks of STS,  none of these studies could be said to incriminate
2,3,7,8-TCDD directly as the agent solely responsible for the excess  of  STS's.
The question remains that the risk of STS may be in part due to the presence of
other polychlorinated dibenzo-p_-dioxin contaminants as well or even to the
phenoxyacetic acid herbicides and/or chlorophenols themselves, or perhaps some
unknown confounder not heretofore discovered.  Hence, based on the epidemiologic
data, the evidence still must be considered inadequate but suggestive that
2,3,7,8-TCDD is carcinogenic to humans.
     With respect to non-Hodgkin's lymphoma (NHL), a similar situation exists.
Data from Hardell et al.  (1981) and more recently Hoar et al, (1986)  support
the  findings of  an excess risk of NHL from exposure to phenoxyacetic acid
herbicides  and/or chlorophenols.  However, this risk is confounded by the
presence of organic compounds also shown to be carcinogenic on their own in the
Hardell et  al. study.   In the Hoar et al. study,  although 2,3,7,8-TCDD-contam-
inated  2,4,5-T was not  identified as being present, other herbicides and insec-
ticides as  well  as  2,4-D were present and were significantly associated with
exposure.   The contaminant 2,3,7,8-TCDD appears not to be present in any of
these  compounds  in  the  Hoar  et al. study although other  isomers of polychlori-
nated  dibenzo-jD-dioxin  were  present. . Wiklund et  al.  (1987) noted a trend of
increasing  risks for  Hodgkin's disease  and NHL with  lapsed time since licensing
of herbicide  applicators  although nonsignificant.  Woods et al.  (1987) reported
significantly increased risks of NHL in  certain occupational subgroups known to
have a high likelihood  of contact with  herbicides.   Buesching and Wollstadt
                                        31                          S'
-------
(1984) found a significant excess of NHL in deceased farmers in Winnebago
County, Illinois, but this study used occupations as listed  on  the  death certifi-
cates as the basis for determining exposure.  No particular  herbicide  or other
chemical was identified however.  Cantor (1982)  noted a similar significant
excess of NHL (reticulum cell sarcoma) in young Wisconsin farmers  in counties
with a high summary measure of agricultural activity, such as number of acres
in small grain acreage, number of acres treated with insecticides,  and number
of acres in wheat acreage.  Again, no particular herbicide,  pesticide, or
chlorophenol was identified.  Burmeister et al.  (1983) also  noted  a significant
excess of NHL in Iowa farmers who were involved  with "egg-laying chickens,
solid milk production, hog products, and herbicide use."  Again, no particular
herbicide was identified.  Similarly, Milham (1982)  noted, in a proportionate
mortality study, a higher nonsignificant risk of NHL in Washington  State paper
and pulp mill workers.  Cantor (1982) stated that 2,4-D and/or  chlorophenols
were chiefly used in these industries.
     Cook et al. (1986), in his cohort mortality study, also reported  a non-
significant excess risk of NHL in chemical  workers exposed to trichlorophenols,
2,4,5-T, and 2,4,5-T formulations.
     Other studies that report slight excesses of NHL or none at all include:
Zack and Suskind, Thiess et al., Lynge, Axelson  et al., Riihimaki et al.,  Fett
et al., Wolfe et al., and Kogan and Clapp.   However, these studies  cannot  be
said to support the argument that there is  no risk because of problems connected
with each of them, i.e., exceptionally small cohorts, insufficient  latency,
insensitivity, and, in some instances, little evidence of exposure.  These
problems were discussed previously.
     The net overall thrust of the data supports the judgment that  there is
limited human epidemiologic evidence that the phenoxyacetic  acid herbicides
                                       32
-------
and/or chlorophenols are causally related to NHL.   However,  with  respect  to
identifying any particular isomer, such as 2,3,7,8-TCDD,  it  must  be  regarded  as
inadequate.
     With respect to stomaqh cancer, Thiess et al.  (1982) and Axel son  et  al.
(1980), in small  cohort studies, found significantly high stomach cancer  risks
based on three cases in each study.  Zack and Suskind (1980), on  the other hand,
did not find any in another small  study.  Only one  other  study produced a
significant excess of stomach cancer, and that is  the case-control  study  by
Burmeister et al. (1983) in which the surrogate for exposure was  occupation as
given on the death certificates.  No other study reported a  signficant excess
of stomach cancer, although most studies have methodological  flaws.   At this
time, the data must be considered only suggestive at best that phenoxyacetic
acid herbicides and/or chlorophenols cause stomach  cancer in humans  based on
the epidemiologic data.  Again, with respect to the potential  carcinogenicity
of 2,3,7,8-TCDD,  the data must be regarded as inadequate.
     Other excess cancer risks by site occur sporatically at best in a few of
these epidemiologic studies.  They must be regarded as inconclusive.  A summary
of the preceding text will be found in tabular form in Tables 2 through 8
following the section on Future Expectations.

                                ONGOING RESEARCH

     Several agencies of the U.S. government are conducting  ongoing  research
that may help to resolve some of the issues raised  regarding interpretation of
the human health studies on 2,3,7,8-TCDD.  NIOSH is conducting both  a morbidity
and mortality study of workers potentially exposed  to TCP, 2,4,5-T,  and penta-
chlorophenol, and therefore 2,3,7,8-TCDD.  Using NIOSH-registry data,  which
                                       33
-------
includes demographic and work history information on some 7,000 employees of 14
different production plants in the United States where  potential exposure to
2,3,7,8-TCDD may have occurred,  NIOSH is  conducting a historic prospective
mortality study.  The results will be reported in the near  future.
     The second major use of the registry data will be  to conduct a hands-on
medical study of some 450 workers at two  of the 14 plants.   Some 450  referents
were age, sex, race, and neighborhood-matched based on  Census block information.
Noninvasive medical tests (including blood specimens) will  be conducted  in
order to evaluate presence or absence of  health conditions  hypothesized  to be
associated with exposure to TCP, 2,4,5-T, and pentachlorophenol  and consequen-
tially to 2,3,7,8-TCDD.  Only the first phase of this study is complete.  That
phase included complete evaluations of approximately  80 cases and  their  refer-
ents.  Two papers are in preparation at this time; they include  an assessment
of chloracne in pentachlorophenol workers as well  as  a  description of medical
results.
     The IARC and National Institute of Environmental Health Sciences are jointly
maintaining an international registry of  persons occupationally  exposed  to
phenoxy acids, chlorophenols, and contaminants during their manufacture  and
use.  This study will pool together cohorts from over 12 countries.   The same
protocol will be used, and the data will  be analyzed  as one big  study.  The
NIOSH registry is included as part of this study.
     The CDC is further conducting pilot  exposure  studies  on former  Vietnam
veterans in conjunction with the Veterans Administration (VA).   Currently,
blood is being drawn from some 600 Vietnam and non-Vietnam veterans and  analyzed
for the levels of 2,3,7,8-TCDD.   This information  will  be  used  to  compare and
validate categories of exposure that have been defined  in  advance  regarding  the
degree of exposure to 2,3,7,8-TCDD sustained by these veterans.

                                       34
-------
     Additionally, CDC and the VA are .corraborating on  a  study of about  40
"heavily" exposed Ranch Hand veterans to determine the  half-life of 2,3,7,8-TCDD
in blood.  Apparently, new blood will be drawn and the  levels  of 2,3,7,8-TCDD
will be determined in that new blood and contrasted with  the levels of 2,3,7,8-
TCDD found in old blood drawn from the same persons at  an earlier time.  This
project was expected to be complete in September of 1987.
     CDC is also planning a series of case-control studies of  liver cancer,
nasal cancer, STS, and lymphomas in Vietnam veterans to be completed in  1989.
     The National Cancer Institute (NCI) is planning a  case-control study of
Nebraska farmers exposed to phenoxyacetic acid herbicides similar to the Hoar
et al. (1986) study of Kansas farmers.  Results will be available in 1989.

                              FUTURE EXPECTATIONS

     The NIOSH studies should provide answers to whether U.S.  workers involved
in the production of herbicides and chlorophenols contaminated with 2,3,7,8-
TCDD suffer from increased risks of site-specific cancer mortality as well  as
other long-term health effects a priori found to be associated with exposure to
2,3,7,8-TCDD.  The CDC studies will help to substantiate or invalidate subjec-
tively defined 2,3,7,8-TCDD exposure gradients in Vietnam veterans as well  as
provide estimates on the half-life of 2,3,7,8-TCDD in blood.  This information
will be  useful in estimating quantitatively the probable dose received by
Vietnam  veterans at the time of exposure to Agent Orange.  A series of case-
control  studies planned by CDC on Vietnam veterans of cancer sites shown
previously to be associated with a higher risk of exposure to Agent Orange  may
provide  additional evidence to either refute or substantiate a higher risk  of
cancer,  especially if it can be determined what the likely dose was at the  time

                                        35                         i .  /
-------
of exposure based on the half-life estimates.   The NCI  case-control study of
Nebraska farmers should serve as a check on the results of the earlier Hoar et
al. (1986) study of Kansas farmers.
                                        36
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-------
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                                      69
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                                                    February 1988
                                                    Review Draft
                      APPENDIX  C
REPRODUCTIVE AND DEVELOPMENTAL TOXICITY OF 2,3,7,8-TCDD
                  G.  L.  Kimmel,  Ph.D.
         Reproductive Effects Assessment Group
    Office of Health  and  Environmental Assessment
          Office of Research and Development
         U.S.  Environmental  Protection Agency
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                               EXECUTIVE SUMMARY

     There is not sufficient evidence to link 2,3,7,8-tetrachlorodibenzo-
p_-dioxin (2,3,7,8-TCDD) to human reproductive or developmental toxicity;
however, it has been shown to be a reproductive and developmental toxicant in
animal studies.  Among the effects that have been reported are reduced
fertility, litter size, postnatal survival, and offspring body weight, as well
as an increase in structural malformations.  Effects on the male and female
gonads and on the female menstrual/estrous cycle have also been reported.
Reproductive and developmental effects have been observed in a variety of
species, indicating that the toxicity is not a species-specific event.
     The studies on reproductive function and fertility remain the basis for
establishing the lowest effective (toxic) exposure level.  It appears that a
0.01 ug/kg/day exposure is the lowest effect level that can be supported by the
data, although further analysis of the data may provide some support for a
lower effect level of 0.001 ug/kg/day.  In addition, a detailed analysis of
studies in the subhuman primate may also provide support for a lower effect
level of 0.001 ug/kg/day.  In relation to developmental toxicity, a large
number of studies in a variety of species has demonstrated that 2,3,7,8-TCDD is
a developmental toxicant.  Collectively, these studies indicate that long-term,
low-dose exposure is of concern relative to the potential for altering
reproductive function and fertility.  The results also demonstrate that acute
and short-term exposures are effective in causing altered development, and
therefore, should also be of concern.
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                                  INTRODUCTION

     2,3,7,8-TCDD has been shown to be a reproductive and developmental
toxicant in animal studies.  Among the effects that have been reported are
reduced fertility, litter size, postnatal survival, and offspring body weight,
as well as an increase in structural malformations.  2,3,7,8-TCDD also affects
the male and female reproductive systems.  Gonadal dysfunction has been
demonstrated in both sexes, and alterations of normal reproductive cycles have
been reported in the female.  Although there have been accidental exposures of
humans to mixtures containing 2,3,7,8-TCDD, there is not sufficient evidence
from the case reports and epidemiologic  studies that have been carried out to
date to link 2,3,7,8-TCDD to human reproductive or developmental toxicity (U.S..
EPA, 1986a).  Much of the information on human exposure  is covered in Appendix
D.
     In line with the original request for the development of this review, this
appendix covers the effects of 2,3,7,8-TCDD on the integrity of  the
reproductive system and  fertility and on prenatal  and  early postnatal
development.  The appendix  is  not inclusive of all studies on these effects.
Rather, it  focuses on the  key  studies and  issues  that  may go  into an  overall
risk characterization.   The present  reviewer  has  tried to present a balanced
view of the studies and  the uncertainties  inherent in  the data  and the
analysis.   Studies on other congeners of polychlorinated dibenzo-p_-dioxins or
of  2,3,7,8-TCDD as a mixture or  contaminant of other agents  are  not included,
except as support where  appropriate.  A  more  comprehensive review of
2,3,7,8-TCDD's toxicity  and its  relation to risk  characterization can be found
in  the Health Assessment Document (HAD)  for Polychlorinated Dibenzo-p_-Dioxins
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(U.S. EPA, 1985).
                                ANIMAL STUDIES
REPRODUCTIVE FUNCTION/FERTILITY-
     The study by Murray et'al. (1979) continues to be the guidepost for
setting standards of exposure relative to reproduction.  The study employs a
multigeneration approach and examines the exposure of male and female rats over
three generations to relatively low levels of 2,3,7,8-TCDD (0, 0.001, 0.01, and
0.1 ug 2,3,7,8-TCDD/kg body weight/day).  The analysis of the data is made
somewhat difficult by considerable variation in the fertility index in both
control and exposed groups.  In addition, the number of impregnated animals in
the exposed groups was lower than desirable (Palmer, 1981).  However, there
were effects that cannot be automatically associated with the variation in the
fertility index, including an increased time between first cohabitation and
delivery, a decrease in litter size, a decrease in the gestational survival
index, and a decrease in postnatal body weight.  Specifically, Murray et al.
reported statistically significant changes in several of the measured
parameters, and these are outlined in Table 1.
     While there is no dispute over the reproductive toxicity seen in this
study, there is some disagreement over the appropriate effect levels.  Murray
et al. (1979) indicated that the lowest statistically significant adverse
effect was observed at 0.01 ug/kg/day and that a no-effect level could be
established at 0.001 ug/kg/day.  In a reanalysis of this study, however, Nisbet
and Paxton (1982) argued that the analysis of Murray et al.  (1979) was limited
by the statistical approach used.  Nisbet and Paxton applied a different
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          TABLE 1.   SUMMARY OF  EFFECTS OF 2,3,7,8-TCDD ON REPRODUCTION
Parameter
Fertility



Litter size



Gestation Survival



Postnatal survival



Postnatal body
weight


Generation
fo
fo
fl
fz
fla
fib

f3
f
fib

fs
fla
fib

?3
fla
fib
•fz

ug TCDD/kg/daya
0.001 0.001
_..
. — —
dec
— dec
• — —
— . — .
— dec
	 dec
._.
-__ .. 	
dec dec
— dec
dec dec
inc —
— dec
— ' —
— —
— , —
— dec
— dec

0.1
dec
dec


dec
dec


decb
	 c


b
	 c


b
	 c


a (—), unaffected; dec, decreased; inc,  increased.
b No liveborn offspring.
c One litter only.
SOURCE:  Murray et al., 1979.
-------
statistical approach which included the pooling of data across all  generations,
and their reanalysis indicated that 0.001 ug/kg/day (the lowest dose used)  was
an effect level and that a no-effect level could not be set. The authors of the
HAD for Polychlorinated Dibenzo-fi-Dioxins (U.S. EPA, 1985) accepted this
argument and used the Nisbet and Paxton reanalysis to establish a lowest
observed adverse effect level (LOAEL) of 0.001 ug/kg/day.  However, the HAD
(U.S. EPA, 1985) also noted that the FIFRA Scientific Advisory Panel (SAP)  did
not feel that the effects were consistent enough at 0.001 ug/kg/day and that
this would have to be considered a no observed effect level (NOEL).  Since this
latter decision by the SAP was made before the Nisbet and Paxton reanalysis, it
is not possible to know if the decision would have been different in light of
the reanalysis.  However, the present reviewer feels that effect levels should
not be set on the basis of the Nisbet and Paxton reanalysis.  While it-appears
that Nisbet and Paxton's approach for increasing the limited statistical power
of the Murray et al. (1979) study is appropriate statistically, it is difficult
to see the biological rationale for pooling the data.  Litters from different
generations (or from subsequent matings within a generation) are not the same.
They have different histories of exposure and each is ti'ed to the effect of the
agent on its parental generation.  Thus, as a general rule, pooling of data
from different generations would seem biologically inappropriate.  Unless some
specific exception can be identified, it is not clear how pooling can be
biologically justified in this case.
     A limited review of a report by Murray et al. (1978), which served as
Exhibit 77 in a 1980 EPA hearing, has raised some questions relative to the
offspring survival.  The Murray et al.  (1979) paper included the standard
parameters of the Gestation Survival Index and Postnatal Survival Index as
-------
measures of offspring  survival  (Table  1).  These parameters showed significant
changes, but not  always  in  a dose-related fashion.  This may be because
offspring viability  is examined during discrete periods of offspring
development and the  investigators do not report viability over the entire early
postnatal period.  Additional data  from the  1978 report by Murray et  al. has
been summarily reviewed  on  the  basis of overall offspring survival, i.e., not
separating the Gestation Survival Index and  Postnatal Survival Index.  There
appears to be a general  pattern of  decreased  survival even at 0.001 ug/kg/day,
if one assumes a  survival rate  of control offspring of 90%.  Appropriate
analytical techniques  would have to be applied to confirm this.  Two  points
should be raised  regarding  the  data.  The first is that the number of offspring
used in the calculations of Murray  et  al. (1978) varied considerably  among the
control and two exposure groups.  How this could affect the parameters is not
entirely clear to this reviewer, but Bailar  (1981) spoke to similar issues in
his testimony and noted  that he found the data suggestive of an effect at the
0.001 ug/kg/day level.   The second  point is  the survival of the control
population of offspring.  In both the fjjj and the f3 litters, survival of the
controls by postnatal  day 21 ranged from 70% to 80%.  Although viability varies
within any laboratory  animal population, this figure seems low and may account
for there not being  an established  decrease  in offspring viability in  these two
groups at 0.001 ug/kg/day.  A more  detailed analysis of this data base may
provide a clearer indication of the potential for decreased postnatal   survival
at the 0.001 level.
     In addition to  the data on offspring survival,  the Murray et al.   (1978)
report summarizes their observations on renal pathology.   When all  observed
effects on the kidney  (i.e., "slightly dilated" and "dilated") are combined,
-------
there appears to be an increase at the 0.001 and 0.01 ug/kg/day in the fja and
f}k litters.  As has been pointed out, it is not entirely appropriate to
combine these two end points, since slight dilation may be due to delayed
development which may be transient in nature.  Nevertheless, the kidney is a
recognized end organ for 2,3,7,8-TCDD effects, and the findings of Moore et al.
(1973) in the mouse indicate that the continuous exposure that is found in a
three-generation study may be more likely to lead to the most obvious effects
on kidney development.  Research in this particular area is continuing.  Abbott
et al. (1987a, b) recently reported that the kidney alterations that occur in
the mouse following a single, prenatal 12 ug/kg dose of 2,3,7,8-TCDD are
consistent with true hydronephrosis.
     Allen and his colleagues examined 2,3,7,8-TCDD effects on reproduction in
the monkey (Allen et al., 1977; Allen et al., 1979; Barsotti et al., 1979;
Schantz et al., 1979).  In a series of studies, female rhesus monkeys were fed
50 or 500 ppt 2,3,7,8-TCDD for up to 9 months.  Menstrual cycles and serum
steroid levels were examined.  Following 7 months of exposure, the females were
bred.  Females exposed to 500 ppt showed obvious clinical signs of 2,3,7,8-TCDD
toxicity and  lost weight throughout the study.  Five of the eight monkeys died
within one year after exposure was initiated.  In a summary of the reproductive
function and  fertility of these animals, Allen et al. (1979) reported that
although the menstrual cycle and menstruation were normal, there was a decrease
in serum estradiol and progesterone in five of eight monkeys.  Only three of
the animals conceived, and only one was able to carry the pregnancy to term.
Females exposed to 50 ppt 2,3,7,8-TCDD in the diet (Schantz et al.,  1979)
showed normal menstrual cycles and serum estradiol and progesterone through 6
months of exposure.  When they were bred at. 7 months, four of eight females did
-------
not conceive and two of four that did could not carry the pregnancies to term.
Only two conceptions resulted in normal births.
     The results of this series of studies could potentially support a lower
LOAEL than that reported by Murray et al.  (1979) in the rat (i.e., 0.01
ug/kg/day).  The high dose (500 ppt) resulted in considerable maternal toxicity
and reproductive dysfunction, while a comparable exposure level  (0.01
ug/kg/day) in the rat did not produce any significant clinical signs in the
parental generation.  This could indicate that the rhesus monkey is more
sensitive to 2,3,7,8-TCDD when exposure occurs over long periods and when
reproductive parameters are the critical end points.  The low dose (50 ppt) was
reported as resulting in specific reproductive dysfunction in the absence of
maternal toxicity.  Since this exposure level is calculated to be approximately
0.002 ug/kg/day, the report suggests that even lower doses are required for
effects on reproductive function than are required in the rat and would support
a lower adverse effect level.  Unfortunately, much of the data on the monkeys
has been presented in abstract form or as part of a review, and consequently, a
critical analysis of the data is impossible.  There has been some indication
that studies of even lower levels (i.e., 5 and 25 ppt) showed signs of
reproductive toxicity, and the data are now beginning to appear in the
literature.  Schantz et al.  (1986) reported altered maternal care of offspring
in monkeys exposed to 5 and 25 ppt for 45 to 49 months, and Bowman et al.
(1987a, b) reported on altered maternal-infant interaction and other
reproductive parameters at the Seventh International Symposium on Chlorinated
Dioxins and Related Compounds.  These reports are now being evaluated and, if
supported, could significantly affect the adverse effect levels calculated for
reproductive and developmental toxicity.
                                      8
-------
     It is important to note that none of these findings establish an
unequivocal effect at the 0.001 ug/kg/day or below level.  However, the
evidence is suggestive enough and the uncertainties are great.enough that it
would seem prudent to consider the 0.001 level as highly suspect.   ,

DEVELOPMENT
     Numerous studies have been done on the developmental toxicity of 2,3,7,8-
TCDD, many of which have been summarized in the HAD for Polychlorinated
Dibenzo-£-Dioxins (U.S. EPA, 1985).  Of particular interest to this review are
those studies which present data that may factor into an overall risk
characterization:  Courtney and Moore (1971), Giavini et al.  (1982a), Khera and
Ruddick (1973), McNulty (1980), Moore et al. (1973), Smith et al. (1976), and
Sparschu et al. (1971).
     Developmental toxicity following exposure to 2,3,7,8-TCDD has been
demonstrated in different species, including the chicken, mouse, rat, rabbit,
ferret, and monkey.  Thus, developmental toxicity of 2,3,7,8-TCDD does not
appear to be related to a species-specific metabolic or physiological response
to exposure.  While specific responses and effective doses do vary among
species and among strains within a species (Courtney and Moore,  1971; Poland
and Glover, 1980), developmental toxicity in response to 2,3,7,8-TCDD exposure
can be expected to occur in all species.
     The exposure range at which developmental toxicity first becomes apparent
is 0.125 to 1.0 ug/kg/day when exposure occurs over a major period of
organogenesis.  The no-effect level appears to be approximately  0.1 ug/kg/day
in rats, mice, and rabbits.  Giavini et al. (1982a) did note an  increase in
extra ribs at 0.1 ug/kg/day in the rabbit.  However, the number  of ribs
                                      9
-------
 normally varies between 12 and 13 in rabbits,  and  it  would  require  a  more
 detailed analysis of the data to establish  this  as a  true effect  level.
 Several  laboratories have examined the effects of  more  acute  exposures  during
 organogenesis.   Moore et al.  (1973)  demonstrated that a single  oral dose of
 1  ug/kg  given  to mice on gestation day 10 produced hydronephrosis.  In  an
 interesting  extension of this finding,  they investigated the  postnatal
 development  of the kidney in  cross-fostered offspring following prenatal
 exposure.  The  frequency of hydronephrosis  seen  postnatally was largely
 dependent  on whether the offspring were nursed by  a dam that  had  also been
 prenatally exposed to 2,3,7,8-TCDD.   Thus,  it  appeared  that continued exposure
 during the lactation period was  required to produce or  maintain the greatest
 effect on  kidney development  during  the postnatal  period studied.
     Except  for the multigeneration  studies, which  tend not to  critically
 evaluate many developmental end  points,  few studies have been carried out on
 developmental periods other than  the  period of organogenesis.   Giavini  et al.
 (1982b) did  examine 2,3,7,8-TCDD  exposure on gestation  days 1-3 in the  rat and
 reported possible  delays  in implantation and some  effect on fetal weight and
 the kidney.  There appear to  be  no studies  on  exposure  during late prenatal
 development.  As  noted  above,  Moore  et  al.  (1973)  examined  effects on the
 kidney postnatally following  prenatal exposure of  the dams,  and demonstrated
 that transfer in  the milk is  a likely contributing  factor to the developmental
 toxicity of  2,3,7,8-TCDD.  There  have also  been  reports  describing the  effects
 of 2,3,7,8-TCDD exposure  on the developing  immune  system (see Appendix  E).
 However, carefully designed studies on  postnatal  exposures  or on changes in
 postnatal  function  following  prenatal exposure have generally not been carried
out, making  it  impossible to  evaluate the potential effect  of 2,3,7,8-TCDD on.
                                      10
-------
the developing young animal.
     In summary, studies on the developmental toxicity of 2,3,7,8-TCDD have
clearly demonstrated that it is a developmental toxicant in a wide variety of
species at relatively low doses.  These studies have also shown that extended
periods of exposure are not necessary for developmental toxicity to result.
Thus, there is a need for concern over acute or short-term exposure to 2,3,7,8-
TCDD.  A composite review and summary of the studies on the developmental
toxicity of 2,3,7,8-TCDD is limited by factors such as the simultaneous
occurrence of maternal toxicity, the use of different animal strains and
exposure routes, and in many studies the small number of animals per treatment
group that are included in the final data analysis.  In addition, there are
differences in study designs and approaches to data analysis which must be
considered in comparing the studies.  These factors do not alter the finding
that 2,3,7,8-TCDD is a developmental toxicant at very low exposure levels.
However, they can potentially affect the final assessment of exposure levels
that can be considered toxic.

MALE AND FEMALE REPRODUCTIVE SYSTEM
     Specific components of the male and female reproductive systems are
affected by 2,3,7,8-TCDD exposure.  In the male, exposures above 1 ug/kg/day
resulted in evidence of testicular atrophy with destruction of the seminiferous
tubules and spermatogenic cells (Kociba et al., 1976;  Norback and Allen,  1973;
McConnell  et al., 1978).  However, following exposures of 0.001 to 0.1
ug/kg/day over a 2-year period,  Kociba et al. (1978) reported that the male
reproductive organs appeared to be unaffected, relative to the controls.   In a
study of offspring of male mice exposed to 0.16 to 2.4 ug/kg/day of 2,3,7,8-
                                     11
-------
TCDD combined with higher'doses of 2,4-dichlorophenoxyacetic acid and
2,4,5-trichlorophenoxyacetic acid, there did not appear to be any effect on
fetal or neonatal development or viability (Lamb et al., 1981).  In the female,
exposure to 1 to 2 ug/kg/day for 13 weeks resulted in changes in estrous
cycles, anovulation, and  signs of ovarian dysfunction (Kociba et al., 1976).
At exposures of 0.001 to  0.01 ug/kg/day in a 2-year study, Kociba et al. (1978)
reported no effects on the female reproductive system.  At 0.1 ug/kg/day, a
decrease in uterine changes such as endometrial hyperplasia were reported.  As
noted previously in the Allen et al. studies, female monkeys exposed to 500 ppt
2,3,7,8-TCDD were reported to exhibit changes in their serum steroid levels.

                                    SUMMARY
     2,3,7,8-TCDD has been shown to be a reproductive and developmental
toxicant in animal studies.  Among the effects that have been reported are
reduced fertility, litter size, postnatal survival, and offspring body weight,
as well as an increase in structural malformations.  Effects on the male and
female gonads and on the female menstrual/estrous cycle have also been
reported.  The effects have been observed in a variety of species, indicating
that the toxicity is not a species-specific event and can be expected to occur
in all species, including the human.
     The studies on reproductive function and fertility remain the basis for
establishing the lowest effective (toxic) exposure level.  There is some
disagreement, centered on the appropriate approach for data analysis, over the
effect level based on the Murray et al. (1979) study.  Based on the current
information from this study, a 0.01 ug/kg/day level is the lowest effect level

                                      12
-------
that can be supported by the data.  However, there is enough suggestive
evidence to indicate a real potential for an effect at 0.001 ug/kg/day.
Further analysis of this study and of the subhuman primate studies by Allen
and his colleagues may provide support for a lower effect level of at least
0.001 ug/kg/day.
     In relation to developmental toxicity, a large number of studies in a
variety of species demonstrated that 2,3,7,8-TCDD is a developmental toxicant.
Although a longer term exposure appears to cause effects at slightly lower
doses, acute and short-term exposures are effective in causing altered
development, and therefore, should also be of concern.  When exposure occurs
during the prenatal period of organogenesis, the lowest effect level is in the
range of 0.125  to 1.0 ug/kg/day and  the no-effect level is approximately 0.1
ug/kg/day.  Studies focusing on other periods of development are limited, but
they do indicate that exposure at any time during prenatal and early postnatal
life must be considered  a  potential  threat to normal development.
     2,3,7,8-TCDD also affects the male and female reproductive systems.
Unfortunately,  the amount  of attention that has been given, to these areas of
investigation has not been as great  as that given to reproductive function and
development.  Gonadal dysfunction has been demonstrated in both sexes, and
alterations of  normal reproductive cycles have been demonstrated in the female.
Continuing investigations  of the effect of 2,3,7,8-TCDD and related agents on
reproductive physiology  and cellular events should increase our understanding
of the potential effect  of 2,3,7,8-TCDD on the male and female reproductive
systems.
     The uncertainties that arise from this data base are many, largely because
the area of reproductive and developmental toxicology covers a wide range of

                                      13
-------
potential exposure-response scenarios, and because the reproductive and
developing systems present constantly changing targets with which a toxicant
may interact.  In the case of 2,3,7,8-TCDD exposure, many of these
uncertainties have been discussed in this review and should be taken into
account in carrying out the final risk characterization.  The laboratory data
establishes 0.01 ug/kg/day as a lowest observed adverse effect level (LOAEL)
when exposure is chronic.  Standard approaches for applying uncertainty/
modifying factors and determining a reference dose (RfD) have been used (U.S.
EPA, 1986b; 1987).  An unequivocal no observed adverse effect level (NOAEL) can
be questioned.  A 1000-fold uncertainty factor (UF) can be applied to account
for variation in sensitivity within the human population, uncertainty in
extrapolating animal data to the human situation, and the use of a LOAEL
instead of a NOAEL in calculating the reference dose.  An additional modifying
factor seems unnecessary, since the uncertainty factor accounts for many of the
concerns of this reviewer, i.e., the true LOAEL/NOAEL and the potential for the
pregnant woman and her offspring to be more sensitive than the average healthy
adult.  An exception to this position would arise if an actual effect level
much below 0.001 ug/kg/day was established, and as noted above, suggestive
evidence is accumulating that would support a lower NOAEL/LOAEL.  The
calculation of a reference dose (RfD) for reproductive and developmental
toxicity, based on the current data and literature base, is as follows:

                                 RfD  = LOAEL/UF
                                      = (0.01 ug/kg/day)/1000
                                      = 1 x 10~5 ug/kg day..-.
                                      14
-------
This does not account for potential or actual human exposures which would have
to be factored into the final risk characterization.
     Future investigations should be encouraged to more clearly define the
substantial data base that already exists.  In the area of reproductive
function and fertility, it would be helpful if a more critical review of the
data of the three-generation rat study and the monkey studies was carried out
to determine if a lower effective dose can be established.  Additionally, a
carefully designed multigeneration study could address some of the limitations
of previous studies and fill certain data gaps.  However, multigeneration
studies are not easily designed, executed, or evaluated, and this step should
only be taken when it is obvious that these data are necessary.  In the area of
developmental toxicology, the most pressing needs seem to be the evaluation of
toxicity during periods of development that have not been adequately assessed,
i.e., the late prenatal and early postnatal periods.  With regard to the
postnatal period, the potential for childhood exposure from breast feeding and
from other environmental sources (e.g., ingestion of soil) seems considerable,
and it has been suggested that the child may be particularly sensitive to
exposure to polychlorinated dibenzo-p_-dioxins.  As relates to reproductive and
developmental toxicity in general, a greater effort should be directed at
identifying the effects of polychlorinated dibenzo-£-dioxins on hormonal
regulation and normal cellular and tissue functions.  There is evidence that
2,3,7,8-TCDD influences steroid metabolism and may be associated with steroid
action at the cellular receptor level.  Pratt and his colleagues (Dencker and
Pratt, 1981; Pratt et al., 1984) have also shown a correlation in various mouse
strains between the susceptibility to induction of cleft palate and the
occurrence of the 2,3,7,8-TCDD receptor, and have proposed that 2,3,7,8-TCDD

                                      15
-------
exerts its teratogenic effect on the palate directly through this receptor.
A considerable amount of literature has developed on the mechanisms of cellular
interaction and action aspects of 2,3,7,8-TCDD, as well as on the
structure-activity relationships of 2,3,7,8-TCDD with other dioxins and related
compounds (recent reviews include: Safe, 1986; Silbergeld and Mattison, 1987).
Efforts should be made to incorporate this information and to evaluate
2,3,7,8-TCDD within the context of the larger family of related agents.  As
additional information becomes available, this review will be updated and
reconsidered relative to its appropriateness in defining critical studies and
issues related to the reproductive and developmental toxicity of 2,3,7,8-TCDD.
                                      16
-------
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                                      18
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Norback, D.H.; Allen, J.R.   (1973)  Biological responses of the  nonhuman
     primate, chicken, and rat to chlorinated dibenzo-p_-dioxin ingestion.
     Environ. Health Perspect. 5:233-240.

Palmer, A.K.  (1981)  Regulatory requirements for reproductive toxicology:
     theory and practice.  In: Kimmel, C.; Buelke-Sam, J., eds.  Developmental
     Toxicology.  New York:  Raven Press.

Poland, A.; Glover, E.   (1980)  2,3,7,8-tetrachlorodibenzo-p_-dioxin:
     segregation of toxicity with the Ah locus.  Mol. Pharmacol. 17:86-94.

Pratt,  R.M.; Grove, R.I.; Kim, C.S.; Dencker, L.; Diewert, V.M.  (1984)
     Mechanism of TCDD-induced cleft palate in the mouse.  In: Poland, A.;
     Kimbrough,  R.D., eds.  Biological mechanisms of dioxin action.  Banbury
     Report No.  18.  Cold Spring Harbor Lab.

Safe, S.H.  (1986)  Comparative toxicology and mechanism of action of
     polychlorinated dibenzo-p_-dioxins and dibenzofurans.  Annu. Rev. .
     Pharmacol.  Toxicol. 26:371-399.

Schantz, S.L.; Barsotti, D.A.; Allen, J.R.  (1979)  lexicological effects
     produced in nonhuman primates chronically exposed to fifty  parts per
     trillion 2,3,7,8-tetrachlorodibenzo-p_-dioxin (TCD.D).  Toxicol. Appl.
     Pharmacol.  48:(2)A180.

Schantz, S.L.; Laughlin, N.K.; Van Valkenberg, H.C.; Bowman, R.E.  (1986)
     Maternal care by rhesus monkeys of infant monkeys exposed to either lead
     or 2,3,7,8-tetrachlorodibenzo-p_-dioxin.  Neurotoxicology 7:637-650.

Silbergeld, E.K.; Mattison, D.R.  (1987)  Experimental and clinical studies on
     the reproductive toxicology of 2,3,7,8-tetrachlorodibenzo-p_-dioxin.  Am.
     J. Ind. Med. 11:131-144.

Smith,  F.A.; Schwetz, B.A.; Nitschke, K.D.  (1976)  Teratogenicity of 2,3,7,8-
     tetrachlorodibenzo-p_-dioxin in CF-1'Mice.  Toxicol. Appl. Pharmacol.
     38:517-523.

Sparschu,  G.L.;  Dunn, F.L.; Rowe, V.K.  (1971)  Study of the teratogenicity of
     2,3,7,8-tetrachlorodibenzo-p_-dioxin in the rat.  Food Cosmet. Toxicol.
     9:405-412.

U.S. Environmental Protection Agency (EPA)  (1985)  Health assessment document
     for polychlorinated dibenzo-p_-dioxins.  Office of Health and Environmental
     Assessment.  EPA/600/8-84/014F.  NTIS PB86-122546/AS.

U.S. Environmental Protection Agency (EPA)  (1986a)   Teratology  and
     reproduction studies with TCDD.  In: Pitot,  H., ed.  The report of the
     "dioxin" Update Committee.  Office of Pesticides and Toxic  Substances,
     Washington, DC.
                                      19
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U.S. Environmental Protection Agency (EPA)  (1986b)
     assessment of suspect developmental toxicants.
     51:34028-34040.
Guidelines for the health
Federal Register
U.S. Environmental Protection Agency (EPA)  (1987)  Reference dose (RfD):
     description and use in health risk assessments.  Integrated Risk
     Information System (IRIS):  Appendix A.  Online.  Intra-Agency Reference
     Dose Workgroup, Office of Health and Environmental Assessment.
     Environmental Criteria and Assessment Office, Cincinnati, OH.
     EPA/600/8-86/032a.
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                                              February 1988
                                              Review Draft
                  APPENDIX  D
    EPIDEMIOLOGIC DATA ON REPRODUCTION AND

          EXPOSURE TO  2,3,7,8-TCDD:

ITS USEFULNESS IN QUANTITATIVE RISK ASSESSMENT
           Sherry G. Selevan, Ph.D.
     Reproductive  Effects Assessment  Group
 Office  of Health  and  Environmental  Assessment
      Office of Research and Development
     U.S. Environmental Protection Agency
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                                  INTRODUCTION                   '

     The following is a discussion of epidemiologic data on the reproductive
effects of 2,3,7,8-tetrachlorodibenzo-p_-dioxin (2,3,7,8-TCDD)  and  their
usefulness in qualitative and quantitative risk assessment.  Several  factors
affect the usefulness of these studies.   First, and probably most  important, is
the assessment of exposure.  Other important factors include the biologic
plausibility of the effects of exposure,  given the time frames of  exposure and
its association with relevant outcomes,  and the difficulty of attributing the
effects of combined exposure to polychlorinated dibenzo-p_-dioxins  (hereafter
referred to as "dioxin"), a contaminant  in herbicides or pesticides.
     Due to the narrow focus of this report, that is, the use of existing
reproductive data in the risk assessment of dioxin, certain restrictions will
be made on the data discussed:  The usefulness of data for risk assessment
depends upon the manner in which the probabilities of exposure for the study
members are determined.  The quality of exposure data may range from very
indirect data to detailed industrial hygiene or environmental  monitoring.
These indirect data are less useful and result from assumptions that
individuals were exposed due to his/her presence in a potentially exposed
region/plant during a specific time period.  More useful data would include
measurement of actual levels of dioxin in air, soil, or water with the most
useful data describing the individuals'  levels of exposure.  Studies with only
indirect, assumed exposure data contribute little to a risk assessment because
of limited confidence in the ability to determine whether a given individual
was actually exposed.  Misclassification of exposure will inevitably occur,
typically resulting in a reduction of the estimate of risk (if such a risk

                                      1
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truly exists).  Studies with more detailed exposure data will contribute more
specific information to such a risk assessment.

                               ROUTES  OF  EXPOSURE
     Humans may be exposed to 2,3,7,8-TCDD through several routes:  dermal,
ingestion, and inhalation.  The rates of absorption through these routes vary,
as do the important routes for each individual.  The routes for each individual
can be affected by such characteristics as location during spraying (e.g.,
Vietnam/Oregon forests), diet (e.g., consumption of contaminated water/fish),
and work practices for those occupationally exposed.  As these characteristics
vary, so will each person's effective dose.  Table 4 (U.S. EPA, 1988) describes
the absorption fraction for dioxin from several routes:  soil ingestion - 0.3,
dermal exposure to soil - 0.005, vapor inhalation - 0.75, fat ingestion from
dairy products or beef - 0.68, dust inhalation - 0.27, fish ingestion - 0.68,
and surface water ingestion - 0.5.  For example, an individual exposed to equal
amounts of 2,3,7,8-TCDD from different routes may have radically different
internal doses (e.g., during the ICMESA plant explosion  in Seveso, Italy, an
individual would have a different potential exposure than.she/he would have if
exposed to soil dust later on).  Consequently, each individual's actual dose  is
difficult to estimate accurately in an epidemiologic study.  Another potential
source of 2,3,7,8-TCDD exposure of potential reproductive importance may be the
infant's exposure through human breast milk (Rappe, 1985; Schecter et al.,
1987; van den Berg et al., 1986).   ,
     The epidemiologic literature, while limited at this point in time, seems
to cover two broad categories:  In occupational settings, paternal effects have
-------
been examined (e.g., aerial sprayers in Vietnam, Vietnam veterans, agricultural
sprayers, and employees of manufacturers of 2,4,5-trichlorophenoxyacetic acid
[2,4,5-T]).  Birth defects and fetal loss have been examined through
environmental exposures (e.g., aerial spraying of 2,4,5-T in Oregon, New
Zealand, and Vietnam, and the ICMESA plant accident in Seveso, Italy).  In
these settings, separation of maternal and paternal effects is very difficult.
The only exception may be in such settings as the plant accident in Seveso
where the woman may already be pregnant and In utero exposures are the ones of
importance.  None of these exposures are "clean;" that is, of dioxih alone.
All the human exposures to dioxin have been as a contaminant of some other
agent (e.g., 2,4,5-T), and typically, only sketchy qualitative and quantitative
exposure data were  available.  Therefore, it is difficult to separate out the
effects of dioxin from the agent it  contaminates.  Only some very general
conclusions  may be  drawn  from the relative toxicities of the associated
compounds.
     A  number  of reviews  have discussed the human data in detail  (Kimbrough et
a!., 1984; Constable and  Hatch,  1985; Friedman, 1984; Hatch,  1984; Hatch and
Stein,  1986; U.S. EPA, 1985).  These efforts will not be repeated here, but the
discussion will be  limited to several key areas of study and  certain  studies  of
special  interest to help  define  the  limits of our  knowledge on the human
reproductive effects of dioxin.

                           STUDIES OF VIETNAM VETERANS

     Two studies have  examined the  reproductive effects of 2,3,7,8-TCDD and
defoliants  in  U.S.  Vietnam veterans.  The primary exposure was to Agent Orange
-------
 which  consists  of defoliants 2,4-dichlorophenoxyacetic  acid (2,4-D)  and
 2,4,5-T,  the latter contaminated with  2,3,7,8-TCDD.   (Agents Purple,  Pink,  and
 Green  were primarily used prior to  July 1965.)   These two  studies  are the  Ranch
 Hand study (Lathrop et al.,  1984) and  the  Center for Disease Control's (CDC)
 study  of  birth  defects (Erickson et al., 1984).   A third study  examines the
 birth  defects in  children born  to Australian  Vietnam veterans  (Donovan et  al.,
 1983).
     The  Ranch  Hand study (Lathrop  et  al.,  1984)  compared  airmen who  had flown
 in the  aerial spraying of defoliants (in fixed-wing  aircraft) to those who
 belonged  to  flying  organizations responsible  for transporting cargo.   Other
 types of  spraying (helicopters  and  backpacks) were not  included in the Ranch
 Hand operations.  Hhile the  data in this report  have not yet been verified
 (through  birth  registration  or  medical  records),  preliminary analyses  have
 examined  various  measures of fertility  and  reproductive success (through
 pregnancy outcomes,  sperm count,  and morphology).  The  researchers are
 currently validating the  interview  data and will  subsequently (they state)  do
 more detailed analyses.   Exposure to 2,3,7,8-TCDD was estimated by calculating
 the amount of 2,3,7,8-TCDD (in  herbicides)  used  during  each  airman's  tour  and
 dividing  this number by the  total number of airmen with the  same
 responsibilities  during the  subject's tour.
     The  Ranch  Hand  study individually matched comparisons to each exposed  man,
 put in  replacements  for refusals, and did not maintain the matches in  the
 analyses.  Due  to these procedures,  determination  of the total study population
 and response rate is  difficult;  1,174 Ranch Handers  were included in the
analysis  of reproductive-data,  as were 1,531 comparisons.   Both the man and his
current or former spouse were interviewed on reproductive history;  research has
-------
shown such interview data on reproductive events.to have greater validity when
collected from the wife.  However, the researchers combined pregnancies from
both spouses (but not necessarily reported by both),  so it is impossible to
examine the women's data separately.  No differences  were observed in early or
late fetal loss, induced abortion, or live births in  a crude analysis of the
two groups.  More detailed analyses examined enlisted men and officers
separately, thus reducing the power of the analyses.   Some important
independent risk factors were not adjusted for in the analysis (e.g., prior
fetal loss in analyses of current fetal loss).  It appears that multiple
pregnancies per family group were all analyzed; however, there was no
discussion of the problems associated with the analysis of nonindependent
events.  A statistically significant excess was found for birth defects,
controlling for parental age and for maternal smoking and drinking; however,
the reproductive outcome of "birth defects" is probably the one which most
needs validation against medical records to assure accurate classification.  No
differences were observed in sperm count or morphology in the two groups.  The
more detailed analysis planned for the future could result in useful
information since the potential for misclassification of exposure in this study
appears to be less than for the other veterans studies.
     Another major study of U.S. Vietnam veterans was a case-referent study of
birth defects done by the CDC (Erickson et al., 1984).  The cases and referents
were drawn from births occurring in metropolitan Atlanta from 1968 through
1980.  The 7,133 eligible cases consisted of live and still births with serious
or major defects identified using the 8th revision of the International
Classification of Diseases (ICD-8).  The referent group was selected from
metropolitan Atlanta live births during the same years; the referents were
-------
frequency-matched on race, year of birth, and hospital of birth.  Power
calculations suggested that a referent group of 3,000 was sufficient; above
that no substantial increase in power would be found for "all defects."  The
authors also stated that power for odds ratios greater than 3.0 for specific
birth defects also would not be substantially affected by using more
comparisons.  To allow for a 70% response rate, 4,246 referent births were
selected.  Seventy percent of the women responded, but only 56% of the men did.
     The potential for exposure for the CDC study was based on three
definitions:  First, a man was considered exposed if he had served in any
capacity at any time in Vietnam before the conception of the infant.  Second,
each veteran was queried about his exposure to Agent Orange.  Third, an
Exposure Opportunity Index (EOI) score was (subjectively) developed by a panel
of specialists who evaluated the veterans' duties, location, and time spent in
Vietnam.  Veteran status was examined to determine whether any military service
might be associated with the 96 birth defects examined.  Veteran status was not
associated and subsequently dropped from further analyses.
     These data were analyzed a number of ways, but for all, the data were
stratified on the three variables used for the frequency matching:  (1) without
potentially confounding characteristics; (2) with key characteristics
identified a priori (maternal age, education, and alcohol consumption, and the
presence of birth defects in close (first-degree) relatives of the child
studied; and (3) with a. posteriori testing of 108 -other characteristics.  In
addition, certain groupings of the 96 birth defect categories, thought to be
related, were also examined.  A limited number of elevated findings were
reported out of almost 400 analyses.  Spina bifida was associated with both
levels of the EOI indexes; cleft lip without cleft palate was associated with
-------
veteran status and high EOI; specified anomalies of nails was reported in
Vietnam veterans; and "other neoplasms" were associated with high EOI.  Due to
the large number of analyses and hypotheses tested, one would expect
approximately 20 significant associations (elevated or decreased odds ratios)
with a significance level of p = 0.05 by chance alone.  Only 11 statistically
significant findings were reported, less than that expected by chance.  The
authors appropriately concluded that "... the data collected contain no
evidence to support the position that Vietnam veterans have had a greater risk
than other men for fathering babies with all types of serious structural birth
defects combined."
     The Australian government examined the association between military
service in Vietnam and birth defects in their offspring (Donovan et al., 1983).
A case-referent study compared births occurring from 1966 through 1979,  of
8,517 children with defects recognizable at birth to an equal number of live
born children without birth defects.  The referents were matched by time of
birth and maternal age.  Exposure was defined as presence of, the father in
Vietnam; no specific index of exposure was developed, but the investigators
stated that such exposure was probably low for the Australian troops in
Vietnam.  No difference lwas observed in exposure patterns between cases and
referents.                                 "          ,
     For veteran studies, the exposures are to the fathers, typically many
years before the pregnancy under study was conceived.  These studies have used
different methods to assign veterans into different exposure groups, based on
service in Vietnam (Erickson et al., 1984; Donovan et al.,  1983), matching of
troop movement with aerial spraying (Erickson et al.,' 1984), or participation
in the Ranch Hand operations (Lathrop et al., 1984).  At the time of these
-------
studies, a suitably sensitive assay for 2,3,7,8-TCDD in serum did not exist.
The CDC recently reported on a new assay they developed in 1986; this assay
yielded serum 2,3,7,8-TCDD levels which correlated well (r = 0.98) with
2,3,7,8-TCDD levels in the adipose tissue of the same individuals (CDC, 1987).
Paired samples collected from the same individuals in 1982 and 1987 were used
to derive the half-life of the 2,3,7,8-TCDD body burden (6 to 10 years).  In
this preliminary report, CDC compared the assayed levels of 2,3,7,8-TCDD to the
exposure categories they had established using the EOI and interview data:  no
association was found between the three methods of scoring for potential
exposure and the serum levels observed.  Furthermore, in their examination of
the sera of Vietnam era veterans (444 Vietnam veterans versus 75 non-Vietnam
veterans), the median sera levels did not differ for these two subgroups.
These data raise a number of issues in the consideration of this body of
research:
     (1)  The relatively long half-life broadens the range of potential
     mechanisms that could occur if an association exists between paternal
     2,3,7,8-TCDD exposure and birth defects.  Friedman (1984) and Hatch and
     Stein (1986) have suggested that such an exposure would cause birth
     defects in offspring through gene or chromosomal mutation of
     spermatogonial stem cells (premeiotic effects).  However, the long half-
     life observed means that postmeiotic effects on sperm are also possible.
     The long half-life also lends more support to Friedman's suggestion of
     maternal/fetal exposures from 2,3,7,8-TCDD in seminal fluid.

     (2)  These data suggest that a great deal of misclassification of exposure
     occurred in the assignment of veteran's to their exposure categories for
     the CDC study.  Plus use of this assay would improve exposure estimation
-------
      in the other  studies.

      (3)   Finally,  the  levels  and  similarity of median 2,3,7,8-TCDD sera values
      in both Vietnam  and  non-Vietnam veterans raises questions  about the
      magnitude of  exposure of  Vietnam veterans.  Other data suggest that the
      exposures of  Vietnam veterans were less than originally thought:  the
      median levels  of 2,3,7,8-TCDD in both groups are below those found in the
      adipose tissue of  the control group of a study in eastern  Missouri
      (Patterson et  al., 1986).

                              OCCUPATIONAL  STUDIES

      Occupational  studies tend to be more useful in the evaluation of a
potentially toxic exposure in  risk assessment.  This is especially true for
qualitative risk assessment, since one can be fairly certain that exposure to
the worker did occur.   If good historical  industrial hygiene data are
available, these can  also be of use in quantitative risk assessment.  As
described in the Vietnam veterans studies,  the exposures evaluated here are
primarily to male workers and the studies have been restricted to examination
of paternal effects;  however, wives may have been exposed indirectly,  through
handling their husband's clothes, etc.   Unlike the veterans studies, the
exposures may be occurring concurrently with conception,  thus increasing the
potential  exposure-  level at the time of the pregnancy.
     Smith et al. (1982) identified 616 male chemical  applicators from a list
maintained by New Zealand's agricultural  Chemicals Board and 531 comparison
workers at small  agricultural contracting companies.  A total  of 89% of the
chemical  applicators  and 83% of the contractual  workers responded to a mailed
                                      9
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questionnaire requesting occupational histories from the men and reproductive
histories from their wives.  Exposure to 2,4,5-T was defined as having sprayed
this pesticide during the year of the pregnancy studied or during the year
preceding that pregnancy.  Group 1 (not exposed to any spraying) consisted of
392 pregnancies; group 2 (sprayed other chemicals than 2,4,5-T) had 109
pregnancies; and group 3 (sprayed 2,4,5-T) had 473 pregnancies.  Comparisons of
group 3 to group 1 found no statistically significant differences in fetal loss
or birth defects.  Other risk factors did not appear to be controlled in the
analysis of these data:  for example, the analysis of miscarriages did not
appear to control for maternal age or occurrence of prior fetal loss.  Multiple
pregnancies per family unit were studied; however, the authors did not address
the problem of nonindependent events.  Additionally, the power of this study
was very low for birth defects.
     Dow Chemical studied  930 male workers potentially exposed to dioxins,
through work with chlorophenol processes, for at least one month from January
1939 through December 1975 (Townsend et al.', 1982).  The reproductive
experience of their  wives,  obtained  by  interview, was compared to the wives  of
an equal number of unexposed male employees.  For the 930 exposed workers, 586
wives were identified and  370 responded  (63%);  for the comparison group,  559
wives were identified and  345 responded  (62%).   Exposure potentials  from  low to
high were assigned to jobs using historic surface contamination data  by an
industrial hygienist familiar with the  history  of the facility.  A pregnancy
was  considered  exposed  if  the employee  had worked in  an  exposed  area  for  at
least  one month  at  any  time prior to conception.  Thus,  this  study,  in  its
analysis, has  similar  problems  to those described above  for the  study of
Vietnam veterans  concerning the time delay  between  exposure and  conception.

                                      10
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The outcomes examined included miscarriages (< 20 weeks gestation),
stillbirths, and birth defects identified (or potentially .identified) before
the child's first birthday.  The relationships of these outcomes were compared
to several classifications of exposure:  any dioxin, only 2,3,7,8-TCDD,
moderate and higher levels of 2,3,7,8-TCDD only, other dioxins only.  The "any
dioxin" and "only 2,3,7,8-TCDD" categories were done as broad groups, plus
split into two groups each, based on duration of exposure (< 12 months, or > 12
months) during the entire employment period preceding conception.  Over two
thousand pregnancies were in the "non-dioxin" group, while the "dioxin" group
consisted of 737 pregnancies.  The "non-dioxin" group included pregnancies
occurring before exposure for exposed workers.  No differences were found in
the two groups.  The power was low for the examination of birth defects.  As
noted above, the assumption of any exposure greater than one month at any time
prior to conception could obscure a true effect in this population.  A
reanalysis of these data, looking at possible associations with paternal
exposure during the 3 to 4 months preceding conception might give more insight
into potential paternal effects associated with this exposure.
     In a clinical, cross-sectional study, reproductive characteristics in
current (N = 131) and retired (N = 161) men who had worked with 2,4,5-T were
selected for comparison to workers without exposure (N = 133).  Of the workers
selected, 55% responded.  Historic exposure data were not available, and job
titles were not sufficient to estimate exposure; therefore,  interview exposure
history was used to group workers into a probable exposure category.
Contamination of the plant with 2,4,5-T occurred in 1949.   For the purpose of
comparing reproductive histories, reported chloracne was used to distinguish
groups.  No differences were found for fetal  loss or birth defects between the

                                      11
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workers with chloracne (N = 107/number of pregnancies = 235) and those without
(N = 91/number of pregnancies = 203) during or after 1948 (when 2,4,5-T was
first handled at the plant under study).  Differences were observed in reports
of decreased libido and difficulty with erection or ejaculation.  Due to the
very indirect nature of exposure assessment, the probability of
misclassification of the exposure category is great.

                       AGRICULTURAL  AND FORESTRY  SPRAYING
     A number of studies have examined the relationship between reproductive
effects and aerial spraying of pesticides (Field and Kerr, 1979; Hanify et al.,
1981; Nelson et al., 1979; Thomas, 1980).  In these studies, exposures could be
to either or both parents.  These are all ecologic studies, in which exposure
of study members is assumed by their presence in an area (e.g., by their
residence), with a specified likelihood of exposure to a given agent.  These
studies are heir to the "ecologic fallacy" in that individuals defined by
residence or some other factor as being in a certain exposure category, may not
in fact belong in that category.  Thus, such studies are of very limited value
in either qualitative or quantitative risk assessments; therefore, only brief
discussions of published reports of these follow.
     The earliest of these studies (Nelson et al., 1979) examined the
association between cleft lip and/or cleft palate (CL/P) in live births
occurring during the 32-year period beginning in 1943 and the spraying of
2,4,5-T in Arkansas.  Approximately 1,200 cases of CL/P Were identified using
both birth certificate data and records of the Crippled Children's Services of
Arkansas Social and Rehabilitative Services.  2,4,5-T was primarily used on
                                      12
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 rice;  therefore exposure was determined by estimating  the  proportion  of  rice
 acreage to total  acreage in each county from data  supplied by  the  Arkansas
 State  Plant Board (1970-1974).   (The exposure definition does  not  address the
 use of 2,4,5-T in forestry.)  The 75 counties were then divided  into  categories
 of low,  medium,  and high potential  for 2,4,5-T exposure.   An increase was
 observed in facial  clefts over time,  which the authors suggested were related
 to improved recording  and case ascertainment and not to an association with
 2,4,5-T spraying.
     A letter  to  the editor of Lancet at approximately the same  time  (Field and
 Kerr,  1979)  discussed  the relationship between 2,4,5-T and 2,3,7,8-TCDD  and
 neural-tube defects in New South Wales,  Australia.  Only limited data were
 presented  in this  letter:   in  New South  Wales,  annual  rates for  neural-tube
 defects  (anencephaly and meningomyelocele)  were compared to annual usage
 figures  for 2,4,5-T for all  of Australia.   Seasonal variation  of 2,4,5-T usage
 was obtained by questionnaire of local governments from the years  1965-1976.
 2,4,5-T  usage  increased  from 90  tonnes  (1  tonne =  1,000 kg) in 1965 to 482
 tonnes  in  1976.  The authors reported  a  linear correlation between the previous
years'  2,4,5-T usage and  neural-tube defects.  In addition, a  seasonal pattern
was noted, with the  highest  rates in conceptions occurring during the summer
months  (December, January,  February).  The  authors also noted that in the
Northern Hemisphere, the  highest  rates in conceptions  also occurred during the
summer months.   In  addition  to the limitations present in ecologic studies,  no
comparison rates of neural-tube defects were reported  in communities without
such exposure,  nor were high versus low exposure communities compared.
     Another letter to the editor of Lancet the next year (Thomas,  1980)
discussed a comparison of  selected birth defects (cleft lip, cleft palate,
                                      13
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spina bifida, anencephalus, and cystic kidney disease) and use of 2,4,5-T in
Hungary.  The birth defects were identified using Hungary's malformation
registry, which records malformations recognized at up to one year of age.  In
Hungary, the use of 2,4,5-T rose from 46 tonnes in 1969 to over 1,200 tonnes in
1975.  The rates of some birth defects (spina bifida and anencephalus)
decreased over the time period examined, while others (cleft lip, cleft palate,
and cystic kidney disease) remained essentially unchanged.  In a comparison of
this report with the New South Wales letter from Field and Kerr (1979), Thomas
noted that:  (1) the population examined in Australia was spread over a much
greater geographic area (thus reducing the probability for exposure), (2) over
half of the population in  Hungary lived in rural areas, while only 15% of the
population in Australia did, and (3) 24.6% of the population in Hungary was
actively involved in forestry and agriculture as compared with only 7.4%  of the
Australian population.  While these data suggest no  association between the
defects  and  2,4,5-T, this  report also did not attempt to compare exposed
communities  with less  exposed  (or unexposed) groups.
     Hanify  et  al.  (1981)  compared rates of diagnosed birth defects  in
stillbirths  (>  28 weeks gestation) and  live births  in Northland, New  Zealand,
to  densities of aerial 2,4,5-T  spray  application  (1960-1977).' The reproductive
events  were  identified through  seven  regional  hospital  records:  3.7,751  births
which  included  436  stillbirths,  264  neonatal  deaths, and  510  with  recorded
birth  defects.   The location,  data,  and quantity of 2,4,5-T  sprayed  was
obtained from  company records,  and monthly estimates were  made  for each  of.the
seven  regions.   Environmental  levels  were  estimated modeling  both  the new
applications plus that fraction of  the  previous applications  thought to  still
be  present.   2,4,5-T was  not sprayed from  1959-1965; thus,  for  this  time
                                • i-''
                                      14
-------
 period,  exposure was considered to be zero.   The zero time period (1960-1965)
 was compared to years thought to be representative of "spraying years"
 (1972-1976).  Associations were found for all  birth defects and for talipes
 under certain assumptions of environmental  persistence of 2,4,5-T.   The authors
 did not  discuss the possibility of secular trends in the data,  due  to other
 medical  or environmental  factors,  or differences in the
 identification/recording  of malformations over time.

                 OTHER ENVIRONMENTAL EXPOSURES -- SEVESO, ITALY

      This  category  of "other environmental exposures"  includes  unplanned
 exposures  such  as exposures  from accidental emissions  from  industrial
 facilities.   With an  immediately recognizable  environmental  incident,  exposure
 of  interest  for  reproduction  could  occur  for both  parents,  or just  for the
 mother,  if she  is already  pregnant  at the time of  first  exposure  (unless there
 could be in  utero exposure to dioxin  in seminal  fluid).   In  1976, during the
 production of trichlorophenol at the  ICMESA plant  in Seveso, Italy, a  runaway
 reaction resulted in  an explosion that ultimately  contaminated 700  acres in the
 surrounding  community.  Environmental levels of  2,3,7,8-TCDD were determined
 using wipe tests, evaluating toxic  effects in  small animals, and analyzing
 grass samples.  Approximately 2 weeks later, over  200 families were evacuated
 from high contamination areas.  Exposures were sufficiently high for chloracne
 to be observed in this environmentally exposed population.  Several  reports
 (Bisanti  et  al., 1979; Homberger et al.,  1979; Pocchiari, 1980; Pocchiari et
 al., 1980; Reggiani,  1978; Rehder et al.,  1978; Tuchmann- Duplessis, 1977)
reported  on comparisons of four potentially affected communities (Seveso, Meda,

                                      15
-------
Cesano, and Desio) to nearby, unexposed commun'ities.  Changes in fetal  loss
rates occurring during the last quarter of 1976 and first quarter of 1977 were
found in both exposed and unexposed communities.  An estimated 150 women were
in the first trimester of pregnancy at the time of the accident (Rehder et al.,
1978); of these, 125 women wished therapeutic abortions by October 1976.
Therapeutic abortions were approved for 30.  Another estimate
(Tuchmann-Duplessis, 1980) reports a total of 108  (50 in 1976 and 58 in 1977)
therapeutic abortions in the four affected communities.  Several reports
(Bisanti et al.,  1979; Pocchiari, 1980; Pocchiari  et al., 1980) suggested  that
a large number  of women obtained unapproved, and therefore not reported,
therapeutic abortions.  This supposition  was supported  by a  steep decrease in
birth  rates in  the  first 6 months of  1977, primarily observed  in the exposed
communities.   (All  communities  had had decreases over time;  however, the
decrease  in exposed communities at this key time was much larger.)  Marked
increases  in  the number of birth defects  were  noted in  1977.   These have  been
attributed to changes  in reporting.   Prior to  this time,  only  certain  birth
defects were  required  to be  reported  to Italian Health  Officers  (Reggiani,
1980).  In addition to  the limitations due to  legal requirements,  there were
other reasons for the  underreporting  of malformations:   "Traditionally, Italian
physicians have under-reported congenital malformations because  of their severe
negative  social implications"  (Tuchmann-Duplessis, 1980a,  b).   Although
physicians and midwives  were encouraged to do  more complete  recording, Reggiani
 (1978) concluded that the  malformation data  were "missing"  for 1976 and
 "incomplete"  for the first trimester of 1977.   In  general,  the researchers
 cited above felt that,  in spite of the problems associated with the data,
 adverse reproductive effects did not occur in this population.  However,  the
                                       16
-------
data do not appear to be complete enough to make definitive conclusions.  In
addition, these reports have the same problems concerning exposure assessment
as those described previously in the Agricultural/Forestry Spraying section.

                           STUDIES OF THE VIETNAMESE

     While studies of the Vietnamese have not been published in Western
journals, a review has been presented by Constable and Hatch (1985).  Two types
of studies have been executed:   In the South, the reproductive experience of
exposed  and unexposed couples was compared; in this case, effects of maternal
and paternal exposures cannot be separated.   In the North, reproductive
experience was compared for families of men who had served in the South  (thus,
assuming paternal exposure) to  families of men remaining  in the North  (and
therefore assuming no exposure  of the father).  Exposures for these studies
have been defined by residence,  using historical data on  spraying and/or
determined by the evidence of destruction of  vegetation.  Thus, these  studies
have the same limitations for use in qualitative and quantitative risk
assessment as discussed previously.
     Only one study  (Lang et al., 1983 described in Constable and Hatch, 1985)
has attempted to determine types of exposure  (e.g., exposed during spraying
versus the more indirect exposure due to dust or diet) and assign exposure
scores.  In this study of veterans in the North, exposure was restricted to the
father at some time  prior to conception, thus resulting  in the same problems
encountered in the American and Australian studies of veterans.
                                      17
-------
                                    SUMMARY
     All three Vietnam veterans studies examined the occurrence of birth
defects.  The Ranch Hand study found an excess for "all birth defects"  (these
were not validated); the CDC study found a difference in exposure for cases
with spina bifida, cleft lip without palate, specified anomalies of the nails,
and "other neoplasms"; the third study (Donovan et.al., 1983) found no
differences.  No associations with other reproductive outcomes were found in
the Ranch Hand study (the only study of the three that looked at the other
outcomes).  No differences were found in reproductive outcome in the three
workplace studies  (Moses et al., 1984; Smith et a!., 1982; Townsend et al.,
1982), however, Moses et al. did report increased impotence and sexual
dysfunction in 2,4,5-T workers with chloracne  (assumed to result from 2,4,5-T
and/or 2,3,7,8-TCDD).  Several of the ecologic studies (Nelson et al.,  1979;
Field and Kerr, 1979; Hanify et al., 1981) reported associations with some
birth defects and  volume of 2,4,5-T application; these data were contested by
Thomas (1980).  Birth defects in Seveso increased after the plant accident,
potentially due to changes in reporting (Reggiarii, 1980).  As presented in this
paragraph, one might conclude that 2,3,7,8-TCDD/2,4,5-T is related to adverse
reproductive outcomes, but as discussed previously, limitations in the design
of the studies neither allow nor rule out such an interpretation.
     Due to the very nature of;the exposures to 2,3,7,8-TCDD, obtaining
suitable data for either a qualitative or quantitative risk assessment is
difficult, if not impossible.  First,  the exposures are primarily in general
environmental  settings (wartime exposures are,  for these purposes,  being
classified as environmental)  where it is difficult to define whether an

                                     18
-------
individual is even exposed and nearly impossible.to quantitate that exposure.
In occupational settings,, both qualitative and quantitative exposure data may
exist, but small population sizes limit the power of the studies.  A nationwide
study planned by the National Institute for Occupational Safety and Health
(NIOSH) on their Dioxin Registry could overcome the problems of limited power.
     All of the studies described above contained a major limitation:
imprecise definitions of exposure of the subjects.  In many studies, the
exposure was assumed by place of residence or presence in a particular area.
Even where more descriptive exposure data are available, such as the CDC and
Ranch Hand studies, misclas.sification may be great (CDC, 1987).  Other studies,
such as the ecologic studies, compared the same area over time rather than to a
comparison area (e.g., Field and Kerr, 1979; Thomas, 1980); thus, differences
in these studies may be due to differences in secular trends rather than to
changes in potential exposure to 2,3,7,8-TCDD.  The evidence from all of these
studies, therefore, is open to question.
     The reports with more detailed exposure information include those
examining paternal exposure.  Kimmel has reviewed the animal literature in
Appendix C and reported that 2,3,7,8-TCDD resulted in testicular atrophy and
reduced sperm count, but .a dominant lethal study found no effects (Kimmel,
1988).  The animal data are too limited, at this time, to suggest the presence
or absence of an effect in the offspring in studies with human male exposure.
The large differences in time frame between male exposure and conception in the
epidemiologic studies probably reduces the chance of identifying any paternally
mediated effect that is potentially occurring.  However, new data on the half-
life of 2,3,7,8-TCDD in humans (CDC, 1987) suggests some exposure may be
occurring many years later,  depending on the level of initial  exposure.
                                      19
-------
Reanalyses of occupational studies, with stricter exposure definitions, might
yield useful information.
     The only studies that address the possibility of female exposure are the
ecologic/environmental studies and the studies of Seveso,  In his review of the
animal'literature, Kimmel (1988) found evidence of reproductive and
developmental effects of dioxin.  The human studies have less informative
exposure data, and so are less useful in drawing conclusions concerning the
reproductive effects of 2,3,7,8-TCDD.  Additionally, in both types of studies,
males may also be exposed, thus clouding the interpretation of such data.
                                      20
-------
                                   REFERENCES
Bisanti, L.; Bonetti, F.; Caramaschi, F.; Del Corno, G.; Favaretti, C.;
     Giambelluca, S.; Marni, E.; Montesarchio, E.; Puccinelli, V.; Remotti, G.;
     Volpato, C.; Zambrelli, E.  (1979)  Experience of the accident of Seveso.
     Proceedings of the 6th European Teratology Conference, Akademiai Kiado.
     Pub.

Centers for Disease Control (CDC).  (1987)  Serum dioxin in Vietnam-era
     veterans - preliminary report.  Morbidity and Mortality Weekly Report.
     36(28):470-475.

Constable, J.D.; Hatch, M.C.   (1985)  Reproductive effects of herbicide
     exposure in Vietnam: recent studies by the Vietnamese and others.
     Teratogenesis Carcinog. Mutagen. 5(4):231-250.

Donovan, J.W.; Adena, M.A.; Rose, G.; Batistutta, D.  (1983)  Case-control
     study of congenital anomalies and Vietnam service.  Canberra:  Australian
     Government Publishing Services.  (As described in Hatch and Stein, 1986).

Erickson, J.D.; Mulinare, J.; McClain, P.W.; Fitch, T.G.; James, L.M.;
     McClearn, A.B.; Adams, M.J.  (1984)  Vietnam veterans' risks for fathering
     babies with birth defects.  JAMA 252(7):903-912.

Field, B.; Kerr, C.  (1979)  Herbicide use and incidence of neural-tube
     defects.  Lancet 1(8130):1341-1342.

Friedman, J.M.  (1984)  Does agent orange cause birth defects? , Teratology
     29:193-221.

Hanify, J.A.; Metcalf, P.; Nobbs, C.L.; Worsley, R.J.  (1981)  Aerial spraying
     of 2,4,5-T and human birth malformations: an epidemiological investiga-
     tion.  Science 212:349-351.

Hatch, M.C.  (1984)  Reproductive effects of the dioxins.  In:  Lowrance, W.W.,
     ed.  Public health risks of the dioxins.  Proceedings of a symposium held
     at The Rockefeller University in New York City, Oct. 19-20, 1983.  Los
     Altos, CA: William Kaufmann.

Hatch, M.C.; Stein, Z.A.  (1986)  Agent orange and risks to reproduction: the
     limits of epidemiology.  Teratogenesis Carcinog. Mutagen. 6(3):185-202.

Homberger, E.; Reggiani, G.; Sambeth, J.; Wipf, H.  .(1979)  The Seveso
     accident: its nature, extent and consequences.  Report from Givaudan
     Research Company Ltd. and F. Hoffmann-La Roche & Co. Ltd.

Kimbrough, R.D.; Falk, H.; Stehr, P.; Fries, G.  (1984)  Health implications of
     2,3,7,8-tetrachlorodibenzodioxin (TCDD) contamination of residential soil.
     J. Toxicol. Environ. Health 14(l):47-93.


                                      21
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Kimmel, G.L.  (1988)  Reproductive and developmental toxicity of 2,3,7,8-TCDD.
     In: A cancer risk-specific dose estimate for 2,3,7,8-TCDD, Appendix C.
     External Review Draft.  U.S. Environmental Protection Agency, Washington,
     DC.

Lathrop, 6.D.; Wolfe, W.H.; Albanese, R.A.; Moynahan, P.M.  (1984)  An
     epidemiologic investigation of health effects in air force personnel
     following exposure to herbicides: baseline morbidity study results.
     USAF School of Aerospace Medicine, Brooks Air Force Base, Texas.

Moses, M.; Lillis, R.; Crow, K.D.; Thornton, J.; Fischbein, A.; Anderson, H.A.;
     Selikoff, I.J.  (1984)  Health status of workers with past exposure to
     2,3,7,8-tetrachlorodibenzo-j)-dioxin in the manufacture of 2,4,5-
     trichlorophenoxyacetic acid: comparison of findings with and without
     chloracne.  Am. J. Ind. Med. 5:161-182.

Nelson, C.J.; Holson, J.F.; Green, H.G.; Gaylor, D.W.  (1979)  Retrospective
     study of the relationship between agricultural use of 2,4,5-T and cleft
     palate occurrence in Arkansas.  Teratology 19:377-384.

Patterson, D.G.; Hoffman, R.E.; Needham, L.L.; Roberts, D.W.; Bagby, J.R.;
     Pirkle, J.L.; Falk, H.; Sampson, E.J.; Houk, V.N.  (1986)  2,3,7,8-Tetra-
     chlorodibenzo-p-dioxin levels in adipose tissues of exposed and control
     persons in Missouri: an interim report.  JAMA 256(19):2683-2686.

Pocchiari, F.  (1980)  Accidental TCDD contamination in Seveso (Italy):
     epidemiological aspects.  FIFRA Docket No. 415, Exhibit 1469.  U.S.
     Environmental Protection Agency, Washington, DC.

Pocchiari, F.; Silano, V.; Zampieri, A.  (1980)  Human health effects from
     accidental release of TCDD at Seveso  (Italy).  FIFRA Docket No. 415,
     Exhibit 1470.  U.S. Environmental Protection Agency, Washington, DC.

Rappe, C.  (1985)  Problems in analysis of PCDDs and PCDFs and presence of
     these compounds in human milk'. , Consultation on organohalogen compounds  in
     human milk and related hazards.  Geneva: World Health Organization.

Reggiani, G.  (1978)  Medical problems raised by the TCDD contamination in
     Seveso, Italy.  Arch. Toxicol. 40:161-188.

Reggiani, G.  (1980)  Direct testimony before the U.S. Environmental Protection
     Agency.  FIFRA Docket No. 415, Exhibit 861.  U.S. EPA, Washington, DC.

Rehder, H.; Sanchioni, L.; Cefis, F.; Gropp, A.  (1978)  Pathological-
     embryological investigations in cases of abortion related to the Seveso
     accident.  Journal of Swiss Medicine  108(42):1617-1625.

Schecter, A.; Ryan, J.J.; Constable, J.D.  (1987)  Polychlorinated dibenzo-£-
     dioxin and poiychlorinated dibenzofuran levels in human breast milk from
     Vietnam compared with cow's milk and human breast milk from the North
     American continent.  Chemosphere 16(8/9):2003-2016.


                                      22
-------
Smith, A.H.; Fisher, D.O.; Pearce, N.; Chapman, C.J.  (1982)  Congenital
     defects and miscarriages among New Zealand 2,4,5-T sprayers.  Arch.
     Environ. Health 37(4):197-200.

Thomas, H.F.  (1980)  2,4,5-T use and congenital malformation rates in Hungary.
     Lancet ii:214-5.

Townsend, J.C.; Bodner, K.M.; Van Peenen, P.P.; Olsen, R.D.; Cook, R.R.  (1982)
     Survey of reproductive events of wives of employees exposed to chlorinated
     dioxins.  Am. J. Epidemiol. 115(5):695-713.

Tuchmann-Duplessis, H.  (1977)  Embryo problems posed by the Seveso accident.
    Le Concours Medical No. 44.

Tuchmann-Duplessis, H.  (1980a)  Direct testimony before the U.S. Environmental
     Protection Agency.  FIFRA Docket No. 415, Exhibit 864.  U.S. EPA,
     Washington, DC.

Tuchmann-Duplessis, H.  (1980b)  Tables in direct testimony before the U.S.
     Environmental Protection Agency.  FIFRA Docket No. 415, Exhibit 864a.
     U.S. EPA, Washington, DC.

U.S. Environmental Protection Agency.  (1985)  Health assessment document for
     polychlorinated dibenzo-p_-dioxins.  Office of Health and Environmental
     Assessment, Washington, DC.  EPA/600/8-84/014F.  NTIS PB86-122546/AS.

U.S. Environmental Protection Agency.  (1988)  Estimating exposures to 2,3,7,8-
     TCDD.  Exposure Assessment Group, Office of Health and Environmental
     Assessment, Washington, DC.  External Review Draft.  EPA/600/6-88/005A.

van den Berg, M.; van der Wielen, F.W.M.; Olle, K.; Van Boxtel, C.J.  (1986)
     The presence of PCDDs and PCDFs in human breast milk from the Netherlands.
     Chemosphere 15(6):693-706.
                                     23
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                                                           March 1988
                                                           Review Draft
                            APPENDIX E
IMMUNOTOXICITY OF 2,3,7,8-TCDD:   REVIEW, ISSUES, AND UNCERTAINTIES
                   Babasaheb R.  Sonawane,  Ph.D.
              Reproductive Effects Assessment Group
          Office of Health and  Environmental  Assessment
                    Ralph  J.  Smialowicz,  Ph.D.
                Health  Effects  Research  Laboratory
                    Office of Health  Research
                     Robert  W.  Luebke,  Ph.D.
                Health  Effects  Research Laboratory
                    Office of Health  Research
               Office  o.f  Research  and  Development
               U.S.  Environmental Protection  Agency
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                               EXECUTIVE SUMMARY    - - .        ,








     2,3,7,8-Tetrachlorodibenzo-jp_-dioxin (2,3,7,8-TCDD)  is a potent immuno-



suppressant in the laboratory animal species studied; however, the immunologi-



cal effects are apparent at exposure levels that also produce other discernible



pathological lesions and reproductive/developmental effects.  There are no



unequivocal cases of significant immune function alterations in humans following



exposure to 2,3,7,8-TCDD.  The cellular and molecular mechanism(s) of 2,3,7,8-



TCDD-induced immunotoxicity is unknown.  Significant data gaps and uncertainties



exist to prevent an immunotoxicity-based health hazard evaluation.








                                  INTRODUCTION








     This document discusses the relevant scientific literature on the immuno-



toxicity of 2,3,7,8-tetrachlorodibenzo-j3-dioxin (2,3,7,8-TCDD) in laboratory



animals and humans.  The document does not provide a comprehensive literature



review of the 2,3,7,8-TCDD-induced immune effects; however, attempts were made



to identify and discuss critical studies and issues, strengths and weaknesses



of the data, and significant uncertainties associated with the evaluation and



interpretation of the data.  Furthermore', significant data gaps in knowledge



are recognized as they may relate to potential risk to the immune system of



humans upon exposure to 2,3,7,8-TCDD.



     The immunotoxic effects of 2,3,7,8-TCDD have been studied by numerous



investigators for over a decade.  Several recent reviews have been published



that provide a very good overview of the animal and human data dealing with



the immune alterations following exposure to 2,3,7,8-TCDD (Dean and Lauer, 1984;




Dean and Kimbrough, 1986; Thomas and Faith, 1985).  The U.S. Environmental





                                       1
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Protection Agency also reviewed the immunoloyical  effects of 2,3,7,8-TCDD  in
the Health Assessment Document for Polychlorinated Dibenzo-j3-Dioxins  (U.S.  EPA,
1985).  The present review updates the immunotoxicity literature on  2,3,7,8-
TCDD and presents issues of concern and uncertainties in  risk assessment.

                                 ANIMAL STUDIES
     Early experimental  work with 2,3,7,8-TCDD revealed that the thymus  is  a
target organ of toxicity.  All animal species that have been studied have con-
sistently displayed involution of the thymus and loss of cortical  thymocytes
following acute or chronic exposure to 2,3,7,8-TCDD, as well as lymphocyte
depletion of T-cell areas of the lymph nodes and the spleen (Harris  et  al.,
1973; Zinkl et al., 1973; Gupta et al., 1973; Vos et al., 1973; 1974;  Luster  et
al., 1979; 1982).  While these changes in the lymphoid tissues were  similar to
those produced by glucocorticosteroids, adrenalectomy failed to prevent  2,3,7,8-
TCDD-induced thymic atrophy or hepatotoxicity (van Logten et al.,  1980).   The
administration of thymosin to mice exposed to 2,3,7,8-TCDD by postnatal  maternal
treatment did not affect thymus atrophy or decrease mitogen responses  (Vos  et
al., 1978).  Thyroidectomy, however, was found to protect rats from  the  immuno-
suppressive effects of 2,3,7,8-TCDD (Pazdernik and Rozman, 1984).   Vos  and
co-workers (1973) reported the effects of 2,3,7,8-TCDD on the immune system of
guinea pigs, rats, and mice.  Several tests were employed in these three species,
and the authors concluded that the effect of 2,3,7,8-TCDD was primarily  on  the
cell-mediated immune function of these animals.  Guinea pigs treated with
1 ug/kg of 2,3,7,8-TCDD per week either died or were moribund after  four doses.
At this dose and also at lower doses (0.008, 0.04, and 0.2 ug/kg), guinea pigs
showed severe loss of body weight, depletion of the lymphoid organs, particularly
-------
the thymus, and lymphopenia.  At sublethal doses, the effects were primarily on
the thymus.  The weights of other lymphoid organs, such as the spleen,  cervical
lymph nodes, and popliteal lymph nodes, were not affected.  Adrenals showed a
slight enlargement at the 0.2 ug/kg dose level.  Serum cortisol  and corticoste-
rone values were not influenced by the 2,3,7,8-TCDD treatment.  No microscopic
lesions were apparent in any of the lymphoid organs upon dosing with up to
0.2 ug/kg of 2,3,7,8-TCDD per week other than atrophy of the thymic cortex (Vos
et al., 1973).
     In selected experiments, the treated and control animals were sensitized
to tetanus toxoid or killed Mycobacterium tuberculosis.  Seven days after the
tetanus toxoid injection, there was a small but significant increase in the
serum, antitoxin levels at 0.008 and 0.04 ug/kg dose levels.  The animals were •
challenged by a second dose of tetanus toxoid 2 weeks after'the first challenge
to evaluate the secondary response.  One and 2 weeks after the second dose, the
guinea pigs treated with 0.2 ug/kg of 2,3,7,8-TCDD per week had significantly
lower tetanus antitoxin titers.  The animals challenged with IM.  tuberculosis
were tested for their skin reactivity 12 and 19 days later by intradermal
tuberculin injections.  Both skin thickness and the diameter of  the reaction
site were significantly decreased, the former being significant at the
0.04 ug/kg dose level (Z.inkl et al., 1973; Vos et al., 1973).
     Female albino rats were treated orally with 0.2, 1, and 5 ug/kg of
2,3,7,8-TCDD once a week for 6 weeks.   The body weights and thymus weights  were
reduced at the highest dose level and adrenal  weights were reduced at the ,1 and
5 ug/kg dose level.  The relative thymus weight (organ/body weight ratio)  was
approximately half that of the control animals at the 5 ug/kg dose level,
whereas at the same dose level, the relative spleen weights were significantly
increased.  The total peripheral blood leukocyte count and lymphocyte numbers
-------
 did  not  show a  significant 2,3,7,8-TCDD-related effect as observed in guinea
 pigs.  Skin  reactivity of rats challenged to tuberculin was not affected at any
 2,3,7,8-TCDD dose  levels (Zinkl et al., 1973; Vos et al., 1973).
     Graft versus  host (GVH) activity of mice was evaluated after 0.2, 1, 5,
 and  25 ug/kg of 2,3,7,8-TCDD treatment per week for 4 weeks.  The absolute
 and  relative thymus weights were decreased in groups treated with 5 ug/kg of
 2,3,7,8-TCDD.  Dose-related reduction in GVH activity was observed (Vos et al.,
 1973).
     Mice treated  orally with 2,3,7,8-TCDD were challenged with either Salmonel-
 la bern or Herpesvirus suis (Thigpen et al., 1975).   In two separate experiments
 the  animals  were exposed to 2,3,7,8-TCDD levels ranging from 0.5 to 20 ug/kg
 once every week for 4 weeks.  The animals were challenged with the infectious
 agents one day after the fourth dose of 2,3,7,8-TCDD.  Treatment of mice with
 2,3,7,8-TCDD reduced the time of death after the bacterial  challenge at the
 5 ug/kg dose level whereas the increased mortality was obvious at 1 ug/kg.  The
 challenge with H.  suis influenced neither the period of time to death nor the
 mortality rate.
     Vos and  co-workers (1978) reported the response of young mice .to
 Escherichia  coli endotoxin after 2,3,7,8-TCDD treatment.   There was a dose-
 related increase in mortality, both with respect to  2,3,7,8-TCDD and the endo-
toxin.   While the administration of 250 ug of endotoxin produced no mortality
 in the control group, as little as 10 ug of endotoxin produced death in mice
treated with 15 or 50 ug 2,3,7,8-TCDD.  A similar increase  in endotoxin sensi-
tivity was reported when mice were treated with a single oral dose of 100 ug/kg
of 2,3,7,8-TCDD and challenged with 20 ug of endotoxin.  In  these experiments,
a 2,3,7,8-TCDD dose-related decrease in body, thymus, and spleen weights was
observed.
-------
     The increased susceptibility to S_.  bern reported in  mice exposed  to
2,3,7,8-TCDD (Thigpen et al.,  1975)  has  been postulated to  be due to an in-
creased sensitivity of 2,3,7,8-TCDD-exposed mice to bacterial endotoxin (Vos
et al., 1978), although variable effects have been  reported in host resistance
following 2,3,7,8-TCDD exposure (Dean and Lauer, 1984;  Vos  et.al.,  1978;  Luster
et al., 1980).  Other studies  have shown a correlation  between increased  mortal-
ity from bacterial or parasitic infections and decreased  serum complement levels
(White  et al., 1986) or decreased B-cell mediated responses (Tucker et al.,  1986)
in 2,3,7,8-TCDD-exposed mice,  respectively.
     Clark et al., (1981) have shown that low doses of 2,3,7,8-TCDD  (4 ng/kg  to
0.4 ug/kg) administered intraperitoneally once a week for 4 weeks  suppress  the
generation of cytotoxic T-cells by lymph node and spleen  cells in male C57BL/6
mice but have little effect on delayed hypersensitivity,  antibody  responses,  or
thymic cellularity.  Suppressor cells capable of blocking the cytotoxic T-cell
response were found in the thymus of 2,3,7,8-TCDO-treated mice following  cumula-
tive doses of 2,3,7,8-TCDD as  low as 4 ng/kg.  The immunosuppressive  effects
observed at 4 ng/kg were significantly less than the 4 ug/kg required  to  induce
thymic atrophy.   In 1983, the  same investigators (Clark et  al., 1983)  reported
that susceptibility to 2,3,7,8-TCDD-induced immunosuppression in mice  is  strain-
dependent, and occurs at doses that have little effect on hepatic microsomal
enzymes.  Clark et al. also reported that other types of haloaromatics can
induce similar immunosuppression provided they possess sufficient  binding
affinity for the  genetically controlled 2,3,7,8-TCDD-receptor protein. These
observations suggest a receptor-dependent mechanism for the stimulation of
suppressor T-cell activity by haloaromatic hydrocarbons.
     Examination  of T-cell-mediated responses in 2,3,7,8-TCDD-exposed  animals
has  generally demonstrated a correlation between thymic atrophy and impaired
-------
cell-mediated immunity.  Exposure of mice, rats, and/or guinea pigs  to
2,3,7,8-TCDD has been reported to result in depressed  delayed  hypersensitivity
(Vos et al., 1973; Faith and Moore, 1977; Clark et al., 1981), prolonged  allo-
graft rejection (Vos and Moore, 1974), depressed GVH response  (Vos and  Moore
1974), decreased responses to T-cell mitogens and/or allogeneic cells  in  vitro
(Dean and Lauer, 1984; Vos and Moore, 1974; Luster et  al.,  1980),  and  decreased
generation of cytotoxic T lymphocytes (CTL) (Clark et  al.,  1981,  1983).
Depressed CTL activity in adult mice following exposure to  2,3,7,8-TCDD has
been reported to be associated with an increase in suppressor  T-cells,  but not
associated with a reduction in the frequency of CTL precursors (Clark  et  al.,
1981; 1983; Nagarkatti et al., 1984).
     Nonspecific immune responses affected by natural  killer cells (Dean  and
Lauer, 1984; Mantovani et al., 1980) and jnacrophages  (Dean  and Lauer,  1984;  Vos
et al., 1978; Mantovani et al., 1980) have not been observed to be affected by
2,3,7,8-TCDD exposure.  Both of these cell types possess tumoricidal ,
bactericidal, and virucidal activity.
     2,3,7,8-TCDD has  also been reported to affect bone marrow and humoral
immune responses of experimental animals.  Exposure of mice to 2,3,7,8-TCDD
resulted  in myelotoxicity and  suppression  of bone marrow progenitor cells
(Tucker et al., 1986;  Luster et al., 1985; Chastain and Pazdernik, 1985).
Studies in mice exposed to 0,  1.0,  5.0, or 15 ug/kg body weight of 2,3,7,8-TCDD
pre- and  postnatally  by maternal dosing  indicated that both 5 and lb ug/kg
dosage groups  had a significant reduction  in bone marrow cellularity,  (colony-
forming  unit-spleen  (CFU-S) or pluripotent stem cells  and colony-forming  unit-
granulocyte/macrophage (CFU-GM).   He'matology  profiles  and blood smears
revealed  a  normocytic  anemia  in; these mice (Luster et  al., 1980).   Bone marrow
toxicity  correlated  with  depressed immunologic  and host-resistance  responses.
-------
 The hematopoietic  stem  cells  have  a  limited  renewal capacity, and damage to



 these  cells  can  induce  a permanent decrease  in their pro!iferative capacity;



 however,  this  limited evidence of  bone marrow toxicity, as observed in mice,



 is  difficult to  evaluate due  to the  lack of  additional significant studies in



 other  species  or humans.



     The  antibody  response of 2,3,7,8-TCDD-exposed mice, as measured by the



 plaque-forming cell  (PFC) assay, was suppressed following immunization with



 T-cell dependent and/or T-cell independent antigens (Tucker et al., 1986;



 Vecchi et al., 1983; Holsapple et al., 1986).  Furthermore, these effects on



 B-cell-mediated  responses were observed to occur at 2,3,7,8-TCDD doses below



 those that cause thyrnic atrophy (van Logten et al., 1980; Tucker et al.,  1986;



 Luster et al., 1985; Vecchi et al.,  1983; Holsapple et al., 1986).



     Susceptibility to suppression of humoral immune responses and 2,3,7,8-TCDD



 induction of the aryl hydrocarbon hydroxylase (AHH) system were found to  corre-



 late (Luster et al., 1980; Tucker et al., 1986;  Chastain and  Pazdernik,  1985;



 Vecchi et al., 1983; Holsapple et al., 1986), as does thymic  atrophy (Poland and



 Glover, 1980), T-cel1-mediated responses (Nagarkatti  et al.,  1984),  and serum



 complement levels (White et al., 1986).  A good  correlation between the degree



 of AHH inducibility and immunosuppression was observed in Fj  crosses of



 "responsive" (Ahb/Ahb)  and "nonresponsive" (AhdAhd)  mice (Nagarkatti et al.,



 1984; Vecchi et al., 1983).  These results, taken  together,  indicate that there



 is a strong association between  the presence of  the  Ah receptor  and  the induc-



tion of immune effects  following 2,3,7,8-TCDD exposure in experimental animals,



and susceptibility  to 2,3,7,8-TCDD-induced immune  effects  in  the  mouse may  be



under genetic influences.



     The immunological  reactivity of young or/suckling rats and  mice has  been




evaluated using a variety of  experimental  protocols .employing prenatal or post-
-------
natal exposure of mothers or a continued prenatal  through postnatal  exposure
with 2,3,7,8-TCDD.  One such study reported by Vos and  Moore (1974)  is  summa-
rized below.  Pregnant rats were treated orally with 1  or 5  ug/kg  body  weight
of 2,3,7,8-TCDD on day 11 and 18 of gestation.  The weights  of  1-day-old  pups
from mothers treated with 5 ug/kg dose level were significantly reduced.   At
this dose level, reduced weights of the thymus and spleen were  also  observed.
Most of the pups in the high-treatment group died within 25  days after  birth.
The newborn animals from mothers given a 1 ug/kg dose level  were further
treated with the same dose of 2,3,7,8-TCDD administered to mothers on day 4,
11, and 18.  Additional groups of suckling rats were exposed to 2,3,7,8-TCDD
postnatally by oral treatment of mothers with 5 ug/kg.   Reduction  in body,
thymus, spleen, and adrenal weights were observed in different  groups at  25
days of age.  Splenic lymphocytes of rats exposed postnatally to the 5  ug/kg
level showed a decrease in phytohemagglutinin (PHA)-induced  DNA synthesis.
This effect was not observed in animals exposed pre- and postnatally to 1
ug/kg of 2,3,7,8-TCDD.  Thymocytes cultured from  postnatally exposed male rats
to 5 ug/kg of 2,3,7,8-TCDD showed a significant reduction of thymidine  incor-
poration in the presence of PHA.  DNA synthesis in response to concanavaltn A
(Con-A), however, was not  reduced in thymocytes.  The authors (Vos and  Moore,
1974) reported the insensitivity of 4-month-old mice to  2,3,7,8-TCDD treatment.
One-month-old mice treated with four weekly doses of 25  ug/kg of  2,3,7,8-TCDD
showed a decreased responsiveness of their splenic lymphocytes to PHA,  whereas
5-month-old animals failed to show this effect  after six weekly doses.
     Similar effects were  observed on GVH activity of spleen cells from 25-day-
old  pre- and/or  postnatally exposed rats.  Only those animals treated with 5
ug/kg  postnatally showed  a  significant  decrease in GVH activity of spleen cells,
Reaction times of heterologous  skin grafts  was  prolonged in  rats  exposed  in
-------
 utero  and  in mice  exposed  pre- and postnatally to 2,3,7,8-TCDD (Vos and Moore,



 1974).



     In  summary, exposure  of mice, rats, and/or guinea pigs (as described



 earlier) to 2,3,7,8-TCDD during the perinatal period resulted in thymic atrophy



 (Faith and Moore,  1977; Vos and Moore, 1974; Luster et al., 1980; Moore and



 Faith, 1976'; Faith et al., 1978).  These animals exhibited a "wasting syndrome"



 that is  similar to that seen in neonatally thymectomized animals that have been



 treated with corticosteroids (Thomas and Faith, 1985).  2,3,7,8-TCDD adminis-



 tered during immune ontogenesis has been observed to affect immune responses



 more dramatically than when administered to adults.   Generally, thymic atrophy,



 suppressed T-cell mediated responses, and bone marrow toxicity have been report-



 ed to be more profound following pre- and/or postnatal exposure to 2,3,7,8-TCDD



 than with adult exposure (Dean and Lauer, 1984; Dean and Kimbrough, 1986;



 Thomas and Faith, 1985; Luster et al., 1979; 1980; Faith and Moore, 1977;  Vos



 and Moore, 1974).  These results suggest that the developing immune system is



 more susceptible to 2,3,7,8-TCDD-induced alterations and, consequently, that



 the very young may be at a higher risk than adults to the immunotoxic effects



 of 2,3,7,8-TCDD.



     Exposure of animals to 2,3,7,8-TCDD has been shown to decrease the respon-



 siveness of lymphocytes to various mitogens in culture.  For example, rabbits



 exposed for 8 weeks to 2,3,7,8-TCDD at 0.01 to 10 ug/kg/week and challenged



 with tetanus toxoid had a decreased PHA response at  the highest dose (Sharma



 et al., 1979).  Sharma et al.  (1979)  also reported that immunologic effects of



 a single dose of 2,3,7,8-TCDD  in mice were  reversible during an 8-week  period



 while the hepatic lesions persisted.   Mice  were treated with a single oral



dose of 10 ug/kg of 2,3,7,8-TCDD,  and selected animals were sacrificed  at  2,  4,



 and 8 weeks after this dosing.   The spontaneous increase of DNA synthesis  in
-------
splenic cultures and a decreased responsiveness of splenic lymphocytes to phy-
tohemagglutinin and pokeweed mitogen was apparent at 2 weeks after the treat-
ment.  The effects persisted up to 4 weeks after the administration of
2,3,7,8-TCDD, but were not noticed when the spleens were obtained from animals
at 8 weeks after the 2,3,7,8-TCDD dosing.  Lymphocyte depletion in the thymus
and reduced thymus weights showed a partial recovery at this time.
     The influence of direct addition of 2,3,7,8-TCDD to mouse-splenic cultures
was reported by Sharma and Gehring (1979).  2,3,7,8-TCDD decreased the unstimu-
1 ated DNA synthesis in lymphocytes at concentrations as low as 10~9 M; however,
no effects were observed in mitogen-stimulated cultures.  Vos and Moore (1974)
reported that the addition of 2,3,7,8-TCDD up to 0.01 ug/0.5 mL in culture
medium did not alter the DNA synthesis in mouse spleen or rat thymus cells,
either with or without the presence of phytohemagglutinin or Con-A.  Luster  and
co-workers (1979) exposed the spleens from mice to 2,3,7,8-TCDD in dimethylsul-
foxide.  DNA, RNA, and protein synthesis in the spleens were inhibited at
2,3,7,8-TCDD concentrations of 10~7 M.  The ability of lymphocytes to bind with
mitogens was not influenced by 2,3,7,8-TCDD.
     The mechanism(s) for 2,3,7,8-TCDD-induced immunosuppression is not fully
known.  However, genetic and structure-activity relationship studies have
provided evidence that 2,3,7,8-TCDD-induced immunosuppression in mice is  asso-
ciated with the presence of the Ah locus and is mediated through a 2,3,7,8-TCDD
cytosol receptor (Nagarkatti et al., 1984; White et al., 1986; Tucker et  al.,
1986; Luster et al., 1985; Vecchi  et al., 1983; Holsapple et al., 1986;  Poland •
and Glover, 1980; Dencker et al.,  1985;  Kouri and Ratrie, 1975).  Some evidence
suggests that thymic epithelial  cells may be the principal  target for 2,3,7,8-
TCDD-induced immunotoxicity (Clark et al., 1983; Greenlee et al., 1985).   Bind-
ing of 2,3,7,8-TCDD to receptors in the thymus may promote altered T-cell  matu-

                                       10
-------
 ration  and differentiation, and it may be the molecular basis for the observed
 thymic  atrophy and immunotoxicity.  Other work, however, suggests that hemato-
 poietic stem cells and B-cells may also be targets for 2,3,7,8-TCDD-mediated
 immune  effects.  For example, it has recently been shown that 2,3,7,8-TCDD selec-
 tively  inhibits the differentiation of B-cells into antibody-secreting cells in
 vitro (Tucker et al., 1986) and inhibits bone marrow stem cell  colony growth in
 vivo and in vitro (Luster et al., 1985).  The binding of 2,3,7,8-TCDD to recep-
 tors on lymphocytes, thymus epithelial cells, and/or hematopoietic precursor
 cells may cause alterations in maturation and differentiation that may result
 in the  immune alterations observed in animals following in vivo exposure to
 2,3,7,8-TCDD,   Further work is clearly needed to investigate the mechanism(s)
 for 2,3,7,8-TCDD-induced immune effects in order to make better estimates of
 the potential  risks associated with exposure of humans  to 2,3,7,8-TCDD.

                                 HUMAN STUDIES             '..'-.

     For a variety of reasons, humans have been exposed to 2,3,7,8-TCDD  in  the
 environment.  The immune function has been examined in  individuals with  pro-
 bable or known exposure to 2,3,7,8-TCDD using assays  designed to evaluate the
 component parts of the immune system.  Several  accounts of immune functions in
 humans exposed to 2,3,7,8-TCDD have been referred to  in summary-type  articles
 reviewed in the preparation of this paper.  Some of these reports have not
 been published in the scientific literature or are anecdotal.   However,  none
 of these reports, in the opinion of the reviewers (Dean and Lauer, 1984;  Dean
 and Kimbrough, 1986; Marshall, 1986), presented convincing evidence for  altered
 immune function in the exposed populations.  In one study, no abnormalities in
measured immune parameters were observed in military  personnel  who had been

                                       11
-------
involved in the spraying of 2,3,7,8-TCDD-contaminated Agent Orange  during
operation Ranch Hand in Vietnam (conversation between J.  Silva,  CHHS,  Test  and
Evaluation Activity, Bethesda, MD, and Dr. Ralph Smialowicz, U.S.  Environmental
Protection Agency, February 6, 1987).
     Attempts have been made to study immune functions in populations  that  have
been exposed to 2,3,7,8-TCDD.  All of these attempts have been complicated  by
technical difficulties in both design and execution, in spite of efforts by
the researchers to control for all possible variables.  Three such studies  are
summarized.
     In July of 1976, an  uncontrolled chemical reaction at a chemical  plant in
Seveso,  Italy, resulted in the release of an estimated 300 g of 2,3,7,8-TCDD
mainly  into an uninhabited area.  However, a significant number of people were
potentially exposed, and  thus, a major health effects surveillance effort was
initiated.  The results of this study, including immunologic assessment, have
been published  (Homberger et  al., 1979; Reggiani, 1980).   A group of 45
2,3,7,8-TCDD-exposed children, 20 of  whom had chloracne, and 44 children without
2,3,7,8-TCDD exposure were evaluated  immunologically  every 4 months for approxi-
mately  1.5 years.  These  studies  revealed no differences between the two groups
in  serum immunoglobulin or complement levels or  in  the ability of their T- and
B-cells to respond to  mitogens in vitro.   Critical  evaluation of these  data are
not possible,  however,  since quantitative data  were not  presented.  It  should
 be noted that  a review (Tognoni  and Bonaccorsi, 1982) of the  Seveso incident,
published 5 years after the  fact, reported increased  serum complement hemolytic
 activity and significantly higher lymphocyte proliferative,responses  in exposed
 subjects.  However, no quantitative data  were presented, the  patient  population
 was not identified per se, and no reference was made as  to how or when  these
 data were collected.  Thus, critical  review and interpretation  of the reported

                                        12
-------
findings  are both impossible and unadvisable.
     The  town of Times Beach, Missouri, was contaminated with 2,3,7,8-TCDD in
1972 and  1973 when waste from a chemical plant was sprayed on roadways to
control dust.  Soil  samples were tested for 2,3,7,8-TCDD 10 years later, and
levels of contamination were so great that the entire town was purchased under
the provisions of the Superfund law (Powell, 1984).  Selected residents were
subsequently classified as having had high or low risk of 2,3,7,8-TCDD exposure
and were evaluated immunologically (Knutsen, 1984).  Tests of delayed type
hypersensitivity (DTH) to a standard battery of skin test antigens, lymphocyte
blastogenic responses to mitogens, and T-cell  subset analysis revealed no sig-
nificant differences between the high- and low-risk groups.  However, there was
a tendency among members of the high-risk group to respond to fewer skin test
antigens and to have slightly different T-cell  subset profiles than low-risk
group members.   Lymphocytes from children'in the high-risk group  also had a
decreased proliferative response to tetanus toxoid compared with  those from
children in the low-risk group.   It should be noted, however, that  in a prelim-
inary study of an unspecified population of Missouri residents exposed to
2,3,7,8-TCDD, no differences were detected in  T-cell  subset profiles, skin test
responses, or lymphocyte proliferation responses (Stehr et al.,  1986).
     Residents  of the Quail  Run  Mobile Home Park in Gray Summit,  Missouri, were
exposed for various  lengths of time to 2,3,7,8-TCDD-contaminated  soil.  Contami-
nation  was the  result of dust control  efforts  using waste oil  containing chemi-
cal, sludge from the  same source as in Times Beach.  Soil sampling studies
revealed levels of 2,3,7,8-TCDD ranging from 39 to 2,200 parts per  billion.
Immunologic testing  was performed on residents  who had lived in the park for
at least 6 months (N  = 154), and their results  were compared to residents
(N = 155) of other mobile parks  in areas where  no evidence of 2,3,7,8-TCDD

                                       13
-------
contamination was found by soil  sampling (Hoffman et al.,  1986).   DTH to  skin
test antigens, antigen- and mitogen-stimulated lymphocyte  proliferation
responses, lymphocyte subset analysis, cytotoxic T-lymphocyte responses,  and
serum immunoglobulin levels were evaluated in both populations,   Some members
of the high-risk group did not respond to skin test antigens  (anergy) and, of
those that did respond, positive reactions were obtained for  fewer antigens
than in the unexposed group (relative anergy).  The exposed group had an  in-
creased frequency of anergy (11.8% vs. 1.1%)  and relative  anergy  (35.3% vs.
11.8%).  It must be pointed out that there were significant technical problems
with the interpretation of the skin test responses and,  as a'result, nearly 50%
of the data had to be discarded.  Several skin test readers,  with various
amounts of experience, were employed in this  investigation.  In addition, two
of the four readers who had received special  training in the interpretation of
DTH skin tests recorded anergy in 15% or 40%  of the control group, a  rate 75 or
200 times the expected rate.  Although allowances were made for'this  by the
investigators, the DTH data in this study are thus questionable.   The mean
ratio of T helper/inducer cells to T suppressor/cytotoxic  cells  was similar  in
both groups, although there was a nonsignificant proportion of exposed  group
members with a T4/T8  ratio less than 1.0  (8.1% vs. 6.4%).   The report  likewise
states that 12.6% of the exposed group versus 8.5% of the  low-risk group  had
abnormal in vitro T-cell functions, although examination of the  tabular data
provides no evidence whatsoever for a difference between levels  of immunocom-
petence in the two groups.  The authors  concluded that long-term exposure to
2,3,7,8-TCDD is associated with depressed cell-mediated immunity, although  the
effects have not resulted  in an excess of clinical illness.  Individuals  from
this initial study  (Hoffman et al,, 1986) have been re-evaluated, and the
results of this follow-up  study, as reported by Evans et al. (1987) at  the

                                       14
-------
International Conference on Dioxin, failed to corroborate the report  of  anergy
in the 2,3,7,8-TCDD exposed cohort of the initial  study (Hoffman  et al.,  1986).

                  UNCERTAINTIES, DATA GAPS, AND RESEARCH NEEDS

     The immune system, unlike many organ systems, is  self-renewing from  a
pool of pluripotent stem cells.  Functional cells are  generally end-stage
cells that have a limited lifetime and are replaced on a regular  basis.
Therefore, unless the store of stem cells is destroyed or unless  there  is a
permanent blockade of cellular differentiation, acute  chemical  exposure  is
unlikely to cause long-term suppression of the host-defense system.   However,
the uncertainty in the premise of self-limited chemical  immunotoxicity  resides
in the unknown effects of chronic or repeated acute, exposure to immunotoxic
agents on the regenerative capacity of the immune system.
     Although a great amount of time and effort has gone into evaluating  the
immunotoxic effects of 2,3,7,8-TCDD in experimental  animals, there,are  a
number of uncertainties that remain to be resolved. These include but  are
not limited to the following:   (1) the lack of a strong association between
2,3,7,8-TCDD exposure and decreased host resistance, (2) the different  suscep-
tibility of species and strains to 2,3,7,8-TCDD-induced immunosuppression,  (3)
the transient and reversible nature of 2,3,7,8-TCDD-induced immune effects,
(4) the apparent but not proven increased susceptibility of very  young  animals
to 2,3,7,8-TCDD-induced immune effects, (5) the uncertainty of the mechanism(s)
of 2,3,7,8-TCDD-induced immune effects, and (6) the lack of evidence  between
observed immunological effects in animals and the questionable immune altera-
tions reported in human populations inadvertantly exposed  to 2,3,7,8-TCDD.
These areas of uncertainty are briefly discussed.

                                       15
-------
     Inconsistent results have been reported regarding the  susceptibility of
2,3,7,8-TCDD-exposed animals to tumor development  and  infection  (Dean and
Lauer, 1984; Faith and Moore, 1977; Luster et al., 1980;  Thigpen et  al., 1975;
White et al., 1986; Tucker et al.,  1986).   Despite the fact that exposure to
2,3,7,8-TCDD results in depressed  T- and B-cell  responses,  further  research is
necessary to define the conditions  under which host resistance is adversely
affected by 2,3,7,8-TCDD.
     In general, the experimental  animal data suggest  that  species differences
exist in 2,3,7,8-TCDD sensitivity.   With the exception of the earliest work,
the mouse has been the predominant species for studying the immunotoxic  effects
of 2,3,7,8-TCDD.  This is most probably due to the fact that more is  known
about the immune system of the mouse than  any other animal  species,  as well as
the fact that there are more validated methods available for examining the
immune system in this species.  Nevertheless, extensive interspecies  comparisons
among several animal species are warranted considering their pharmacokinetic
differences.  These studies are needed not only to substantiate  and  corroborate
the results of studies with the mouse but  also to  provide the framework  to
extrapolate the potential for 2,3,7,8-TCDD to affect the human  immune system.
Interspecies studies are also necessary in order to determine if an  association
exists between the Ah locus and 2,3,7,8-TCDD-induced immune effects  in  species
other than the mouse.  This includes extension of  in vitro  studies which have
demonstrated the presence and inducibility of AHH  in human  lymphoid  tissue
(Kouri and Ratrie, 1975).  Hopefully, these studies will  provide useful  infor-
mation about the role that the Ah  locus may play in the susceptibility  of the
human population to 2,3,7,8-TCDD.
     The immune effects that have  been observed following 2,3,7,8-TCDD  exposure
have, in all cases where it has been examined, returned to  normal levels over

                                       16
-------
a period of time after the cessation of exposure to 2,3,7,8-TCDD.   This  is
probably a result of the plastic nature of the immune system, >as wel 1  as  the
fact that chronic dosing with 2,3,7,8-TCDD has not been performed.   It is not
clear whether and how repeated doses might affect the immune, system or whether
short-term exposure could result in irreversible effects.   Low-dose chronic
studies in animals are needed to determine whether such exposures not  only
produce immune alterations but also to determine if long-lasting  impairment is
produced.  Low-dose chronic studies are important to extrapolate data  from  the
high dosages used under experimental situations because they are most  likely
to mimic the form of human exposure to 2,3,7,8-TCDD in the  environment.
     Exposure to 2,3,7,8-TCDD during immune system development  has  been  shown
to affect the immune system of experimental  animals (Faith  and  Moore,  1977;
Vos and Moore, 1974; Luster et al., 1980; Faith et al., 1978).  The effects
produced following perinatal  exposure have been reported to be  more profound
than those produced following adult exposure, although in  some  cases a high
degree of fetal  toxicity has  been reported (Vos and Moore,  1974).   While  this
may be true, no attempt has been made to test this hypothesis  by making  a
direct comparison between perinatal and adult exposure to  2,3,,7,8-TCDD using
identical exposure regimens (i.e., B- and T-cell function,  host resistance
models, natural  killer cell  activity, etc.).   Work is warranted in  this  area
in order to provide evidence of an increased risk of the developing immune
system to 2,3,7,8-TCDD exposure.  This is necessary so that an  informed judg-
ment can be made as to the potential increased relative risk to infants  and
children exposed to 2,3,7,8-TCDD.
     There is still a great deal of uncertainty about the mechanism(s) by which
2,3,7,8-TCDD causes immune alterations in animals.  Work with other animal
species and in vitro work with human tissues  will hopefully provide new  insights

                                       17
-------
In this area.  A clearer understanding of the mechanisms  of 2,3,7,8-TCDD-
induced immunosuppression in animals will  be invaluable for extrapolation of
the potential health risk in humans.
     Laboratory studies have clearly demonstrated the immunotoxic  effects of
2,3,7,8-TCDD in a variety of animal  models.   Extrapolation  of  these  data to
predict effects of human 2,3,7,8-TCDD exposure are complicated at  best, since
actual exposure levels, route of exposure, and even comparability  of human and
animal susceptibility to the toxic effects of 2,3,7,8-TCDD  are unknown.  Results
of rodent studies suggest that the immune responses of children  should be more
sensitive to the immunotoxic effects of 2,3,7,8-TCDD than that of adults (Faith
and Moore, 1978).  However, immune function  was followed  in children from the
most heavily contaminated area of Seveso for 18 months after the accident and
appeared to be normal (Homberger et al., 1979; Reggiani,  1980).   Furthermore,
baseline data for the immune system of children is not readily available, and
the normal response in children of various ages is not well defined.  Further-
more, altered immunocompetence has been reported in 2,3,7,8-TCDD-exposed
residents of Missouri (Knutsen 1984; Hoffman et al., 1986), although the differ-
ences between mean values for control and exposed populations were not statis-
tically significant.  Suppression was not detected in functional  parameters;
rather, trends in the distribution of exposed group members into "normal" and
"abnormal" response categories were cited as indicative  of  immune dysfunction.
While these trends may or may not be related to 2,3,7,8-TCDD exposure, there  has
been no report of an increase in clinical illness attributable to suppressed
immune function.  Thus, it appears to be that there are no  unequivocal  cases  of
significant immunotoxicity in humans following 2,3,7,8-TCDD exposure (Dean and
Lauer, 1984; Dean and Kimbrough, 1986; Marshall, 1986; Evans et al., 1987).
                                       18
-------
     The  available animal data reviewed above suggest that 2,3,7,8-TCDD-induced
thymus atrophy and immune alterations may result from direct actions on peri-
pheral lymphocytes or progenitor lymphoid cells in the bone marrow and through
altered differentiation of intrathymic precursor cells, specifically by a direct
action on thymic epithelium.  The induction of T-cells, cytotoxic for tumor
target cells, was found to be impaired in mouse studies conducted by Clark et
al. (1981) following total exposure to a 2,3,7,8-TCDD dose of 4 ng/kg over a
4-week period.  These dosages are below those levels significantly altering
other cell-mediated immune responses (Clark et al., 1981); however, these
results are unconfirmed and questionable.  For example, the kinetics of CTL
suppression was not defined in the Clark et al. studies (1981,  1983), and the
effect of 2,3,7,8-TCDD on CTL-mediated tumor resistance remains to be resolved.
Additional concern is raised about Clark's findings (Clark et al., 1981) by
researchers at the Chemical  Industry Institute of  Toxicology, (CUT)  who claim
that the CTL response was not suppressed at a dose of 4 ng/kg (conversation
between Jack Dean, CUT, Research Triangle Park, NC, and Bob Sonawane,  U.S.
Environmental Protection Agency,  February 17, 1987).  The £059  (dose producing
50% maximal  response)  for the induction of thymic  atrophy in sensitive  mouse
strains is approximately 10 umol/kg  (Poland and Glover, 1980) and for antibody
plaque-forming cell  (PFC) suppression varies by 30-fold (1 ug/kg to 30  ug/kg)
between C57BL/6 and DBA/2 mice (Dean and Lauer, 1984).   Furthermore, the issue
of host-resistance effects following exposure to 2,3,7,8-TCDD is unresolved.
     In summary, it may be inappropriate to derive an immunotoxicity-based
hazard assessment for 2,3,7,8-TCDD from mostly acute and/or subchronic  type
studies.   A three-generation  reproductive study by  Murray et  al. (1979)  demon-
strated the critical  noncancer end point for adverse effects at 1 ng/kg/day
compared  to questionable immunosuppressive effects  observed in  mice  by  Clark

                                       19
-------
et al. (1981) at 4 ng/kg by parental  administration.  It seems that the level
of concern identified in either of these studies or in  chronic bioassays  for
carcinogenicity is essentially the same.  The magnitude of differences  rein-
forces a common conclusion that the biological  significance of any  immunological
effects observed in laboratory animals is not adequately established to support
its use as the critical  end point in  hazard evaluation  of 2,3,7,8-TCDD  exposure
to humans.
                                       20
-------
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Vos, J.G.; Moore, J.A.; Zinkl, J.G.   (1974)  Toxicity of 2,3,7,8-tetrachloro-
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Vos, J.G.;  Kreeftenberg,  J.G.; Engel, H.W.; Miderhoud, A.; van  Noorle, J.L.M.
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Zinkl, J.G.;  Vos,  J.G.;  Moore, J.A-.;  Gupta,  B.N.   (1973)   Hematologic and
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      Environ.  Health  Perspect. 5:111-118.
                                        24
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                                                 June 1988
                                                 Review Draft
                    APPENDIX  F
       RATIONALE  FOR A  HORMONE-LIKE MECHANISM

    OF 2,3,7,8-TCDD FOR  USE  IN  RISK ASSESSMENT
                 Michael A.  Gallo
    Professor and Chief, Division of Toxicology
Department of Environmental and Commmunity Medicine
University of Medicine  and Dentistry  of  New  Jersey
         Robert Wood Johnson  Medical School
                   675 Hoes Lane
           Piscataway,  NJ  08854-5635
   Office of Health and Environmental Assessment
         Office of  Research  and Development
       U.S.  Environmental Protection Agency
-------
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     The mechanisms of action of 2,3,7,8-tetrachlorodibenzo-|)-dioxin (2,3,7,8-
TCDD) have been intensely investigated since the pioneering work of Kimmig and
Schultz (1957) in elucidating the chloracnegen in the chlorophenol processes.
The most complete and thought-provoking treatise on the subject is that of
Poland and Knutson (1982).  These authors and others in Poland's laboratory
drew upon their work and others to present a unified hypothesis to account for
the varied responses in animals exposed to 2,3,7,8-TCDD.  In essence, this
hypothesis, which is partly based on the classic receptor theories invoked for
steroid action, suggests that (1) there is a cytosolic receptor for
arylhydrocarbons (the Ah receptor) that binds several compounds and then
translocates to the nucleus,' and (2) there is a second stage to the toxic
reaction(s) that is related to, but not congruent with, the induction of
cytochrome Pi-450.  Activation of this receptor leads to a cascade of reactions
culminating with the association of the receptor-2,3,7,8-TCDD complex with
nuclear DNA.  This association leads to the synthesis of a specific microsomal
protein designated as cytochrome Pj-450.  To date, the chemical that binds this
putative receptor with the greatest avidity is 2,3,7,8-TCDD.  However, many
other halogenated and non-halogenated compounds also bind to this cytosolic
protein (Nebert et a!., 1972).  The xenobiotics with the greatest affinity are
those that are planar, with two phenyl rings and contain substitutions in the
lateral positions.  Several investigative teams have examined the structure
activity relationships (SARs) among the polyhalogenated biphenyls (PCBs, PBBs),
polychlorinated dibenzo-j)-dioxins (PCDDs), and polychlorinated dibenzofurans
(PCDFs) (Knutson and Poland, 1980).  Safe and his co-workers have synthesized
several of the highly active PCBs and have demonstrated remarkable SARs for
several biological end points (Mason et al.,  1987).  Excellent reviews on the
                                      1
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SAR for PCDDs, PCDFs, and PCBs can be found in recent issues of the Annual
Reviews in Pharmacology and Toxicology (Vols. 22 [Poland and Knutson, 1982] and
26 [Safe, 1986], respectively).  The hypothesis also states that ".  . .there is
a second stage to the toxic reactions of 2,3,7,8-TCDD that are related to, but
not congruent with, the induction of cytochrome Pj-450."  This portion of the
hypothesis is supported by the reports of Poland and his co-workers with XB
cells in culture (Knutson and Poland, 1980), Safe and co-workers' findings of
PCB inhibition of 2,3,7,8-TCDD toxicity (Haake et al., 1987), and the Umbreit
et al. report of apparent maximal induction of arylhydrocarbon hydroxylase (the
major system affected by cytochrome Pj-450) (Umbreit et al., 1987) by complex
mixtures of polyaromatic hydrocarbons and very low bioavailability of
PCDDs/PCDFs as measured by tissue levels and signs of toxicity.
     At this point in time, it appears clear that 2,3,7,8-TCDD is working
through the proposed receptor mechanism for the first phase of its activity
(i.e., binding to a putative cytosolic receptor with subsequent  induction of
P-450).  However, all of the biological effects of this molecule cannot be
explained by simple receptor binding and induction of cytochrome PI-450.  The
recent evidence from several laboratories has expanded on the  initial studies
of Poland and Knutson (1982), Neal et al.  (1982), and Barsotti et  al. (1979) to
show quite dramatically that 2,3,7,8-TCDD markedly affects the interaction of
steroids with their respective receptors (Romkes et al.,  1987; Gallo et al.,
1986) and 2,3,7,8-TCDD alters the number of Epidermal Growth Factor  (EGF)
receptors in susceptible cell lines  (Matsumura et al., 1984).  Molloy et  al.
(1987) recently reported the alteration of specific epidermal  keratins in the
HRJ/S strain of mice after treatment with 2,3,7,8-TCDD.  This  finding is
especially relevant to the database since it was in this strain  of mice that
-------
 Poland  and Knutson  (1982)  reported the model for chloracne.  Matsumura et al.
 (1984)  studied the  role of 2,3,7,8-TCDD and EGF receptors, while several
 laboratories  (Safe,  1986;  Gierthy et al., 1987; Umbreit and Gallo, 1988;
 Goldstein et  al., 1987) have been pursuing the interactions of 2,3,7,8-TCDD
 with steroids, primarily estrogen-sensitive tissues.  The interactions with
 glucocorticoids  has  been studied extensively by several laboratories  (Luster et
 al., 1984; Sunihara  et al., 1987).  In general, the response can be summarized
 as a decrease in the number of available cytosolic receptors for EGF or the
 steroids without a decrease in the affinity for the respective ligand.  This
 phenomenon is termed "down-regulation" of the cytosolic receptor.  The
 measurement of receptor binding is biochemical.  The estimate of affinity and
 binding site number  is by extrapolation of the response curve(s)  by Scatchard
 analysis (1949).   The strengths of this analysis are obvious, but the
weaknesses are difficult to reconcile.  The major weakness is the lack of
direct binding information; this does not allow segregation of the receptors by
tertiary structure nor does it completely account for nonspecific binding.  The
 second weakness of ligand binding experiments is the inability of the analysis
to shed any light on the reason for the down-regulation.  It must be emphasized
 at this point that none of the steroid receptor research has demonstrated an
 antagonism between 2,3,7,8-TCDD and the endogenous steroid for the respective
 steroid receptor, nor has any competitive binding of steroids by the Ah
receptor been demonstrated.  However,  the steroid receptors are products of a
supergene family which is responsible for the protein synthesis of all these
receptors (Nebert et al.,  1972),  and the  Ah receptor has many of the structural
and functional characteristics of the steroid receptors (Poellinger et al.,
1986).
-------
     To better understand the role of 2,3,7,8-TCDD in cellular function (or
dysfunction), one must look to the results of the laboratories working on the
mechanisms of action of 2,3,7,8-TCDD at the molecular level.  The major groups
involved in this research are Poland, Nebert's group at the National Institutes
of Health, and Whitlock's laboratory at Stanford University.  As stated above,
Poland established the role of the Ah receptor in some of the actions of
2,3,7,8-TCDD.  Whitlock (1987) has summarized his data and that of other
investigators regarding the regulation of the cytochrome P-450 gene family,
along with the data supporting the hypothesis that the Ah locus is part of a
super gene family responsible for the metabolism of xenobiotics and endogenous
compounds.  Recent work in this laboratory has also elucidated a region on DNA,
which is sensitive to the 2,3,7,8-TCDD-cytosolic receptor complex, upstream
from the cytochrome Pj-450 gene (Neuhold et al., 1986).  These findings are
critical in light of the findings of the 2,3,7,8-TCDD-  responsive gene
expression enhancer system (region) described by Whitlock (Jones et al:, 1986).
Hence, the two laboratories have defined the regulatory mechanisms by which
2,3,7,8-TCDD controls gene expression of the cytochrome Pj-450 (Whitlock,
1987).  The significance of these findings for 2,3,7,8-TCDD risk analysis is
the congruency between gene regulation for the Ah receptor, the glucocorticoid
receptor, and the estrogen receptor (Becker et al., 1986).  The importance of
these findings cannot be underestimated.  There is direct analogy with the
steroid receptor mechanisms and the control of the steroid receptor messenger
RNA (Yamamoto, 1985).  The role of the estrogen receptor (ER), and other
steroid receptors, is understood to a greater extent than the Ah receptor
probably because of the greater emphasis on the physiology of steroids.  The
analogy between the receptor complexes and DNA leads to the obvious comparison
-------
 of effects  of 2,3,7,8-TCDD and  steroids.   Many  of  the  changes  seen  in  animals
 after dosing with 2,3,7,8-TCDD  mimic  estrogen or antiestrogen  effects.   Umbreit
 and Gallo  (1988)  reviewed  these findings,  which are  presented  in Table 1.
 Kociba et  al.  (1978)  demonstrated  the hepatocarcinogenic  effect of  orally
 administered 2,3,7,8-TCDD,  but  in  the same study there was  a marked
 dose-dependent decrease  in  tumors  of  the mammary glands and uteri which  are
 estrogen-sensitive organs.   These  are highly significant  observations  which
 have been pursued by  some  laboratories.  If 2,3,7,8-TCDD  is acting  through
 hormonal  (estrogen) mechanisms,  then  alteration of ovarian  function, exogenous
 estrogens,  or  antiestrogens  should modify  the response(s) to 2,3,7,8-TCDD.
 Recent results have shown  that  2,3,7,8-TCDD effects  can be  overridden  by
 exogenous estradiol (Gallo et al., 1986) and the down-regulation of the
 estrogen receptor is  also  antagonized  by estradiol (Romkes  et  al.,  1987).  The
 significance of these findings  are amplified if one  couples the reports of
 regulation  of  the EGF receptor  by estrogens (Mukku and Stance!, 1985; Madhukar
 et  al., 1984)  along with the consistent observation  that the lowest doses in
 the  lifetime bioassays of 2,3,7,8-TCDD decrease tumor yield in rodent livers
 but do not  affect the backround levels of  breast or  uterine tumors  (Kociba et
 al., 1978).  2,3,7,8-TCDD inhibition of tumor growth at low doses and
 enhancement at higher levels (in the bioassays)  is supported by the recent
 report of a marked decrease in tumorigenesis in  the two-stage liver model at
the lowest dose of 2,3,7,8-TCDD after diethylnitrosamine (DEN)  initiation
 (Pitot et al., 1987).   These findings are consistent with the hypothesis that
2,3,7,8-TCDD may  be working through an endocrine-sensitive mechanism to yield
its toxic effects.  If one accepts this premise, then it is reasonable to
assume that the actions of 2,3,7,8-TCDD can be  explained using  a
-------
         TABLE 1.   ASSOCIATION OF EXTROGENS WITH  2,3,7,8-TCDD  TOXICITY

    Many of the toxic effects of 2,3,7,8-TCDD are similar to effects  of
    estrogens in non-2,3,7,8-TCDD treated animals.

1.   Some effects of 2,3,7,8-TCDD resemble effects of elevated
    estrogens.

2.   Other effect of 2,3,7,8-TCDD resemble antiestrogenic effects
    [most antiestrogens have estrogenic effects at different  doses].

3.   Some 2,3,7,8-TCDD effects are not straightforward estrogenic or
    antiestrogenic effects.  For some of these, an influence  of
    estrogen on the effect is known.

4.   Other signs of 2,3,7,8-TCDD toxicity may be related to cholesterol
    mobilization for estrogen synthesis.
-------
                              TABLE  1.   (continued)
                              2,3,7,8-TCDD EFFECTS
 Fat  loss  1,4
 Wasting  1,4
 Changes  in serum  lipids:  1,3,4
   increased cholesterol
   increased LDL,  VLDL
 Anorexia  1
 Hypophagia 1
 Hypoinsulinemia 1
 Altered serum  fatty  acids  1,4
 Hypoglycemia 1,4
 Lowered Q£ consumption
Immunosuppression 1
Thymic involution 1
Decreased thymic
  cellularity
Hirsutism 1
Chloracne 3
Skin keratinization 1
Membrane damage  1,4
Stimulates differentiation 1
   in certain epithelial
   cells
Lowered T4 in serum 3
Increased:
  thyroid weight 3
  serum TSH
  T4 excretion as
    glucuronide
Lower serum testosterone 3
Uterine suppression 2
Reproductive failure 1,2,3
Blockage of E2 uterotrophism 1
Terata 1,2,3
-------
                             TABLE 1.  (continued)

Lowered serum corticoids 3,4
Blockage of ACTH stimulation of corticosteroid synthesis 3,4
Downregulates:
  E6F receptor 1
  Prolactin receptor 1
  Glucocorticoid receptor 1
  LDL receptor 1
  Estrogen receptor 1

Ascites 1,4
Hepatocyte membrane damage 1,4
Hepatocyte membrane cAMP reduced
Enzyme inductions 3
  AHH (EROD, P-448, P-450c, and d)
  Ornithine decarboxylase
  UDP-GTs
  ALA synthetase
  Others
Anemia
Porphyria cutanea tarda 1
Iron accumulation in liver
Altered iron transport in gut
                                      8
-------
physiologically based model.  The physiological implications of an endocrine
mechanism can explain many of the responses seen in animals after exposure to
2,3,7,8-TCDD (see Table 1) since these responses are similar to hyper- and
hypo-hormonal states (O'Malley and Buller, 1977; Potter et al., 1986; Jones et
al., 1986; Gustafsson et al., 1987).  As stated above, the analogy to the
estrogen system is arguably the strongest to 2,3,7,8-TCDD effects (both
hyperplastic and dysplastic responses in endocrine sensitive organs), and the
similarity between the cytosolic receptors and their stabilization by molybdate
(Denison et al., 1986), activation of a nuclear site, and the anti-2,3,7,8-TCDD
effect of estradiol strengthens the analogy.  The effects of estrogens are
widespread throughout the body.  Some of these effects may not be receptor-
mediated, but the majority of the effects are directly attributable to receptor
binding.  The toxic effects of estrogens have recently been summarized (Umbreit
and Gallo, 1988) and include thymic involution and decreased response to septic
challenge (Luster et al., 1984; Grossman, 1984), a wasting syndrome
characterized by weight loss, hirsutism, and epidermal lesions.  As a note,
there are recent reports of estrogens enhancing or causing cachexia and wasting
which are major effects of 2,3,7,8-TCDD seen in intoxicated animals.  The role
of estrogens as immunosuppressives is not well understood, but it is
hypothesized that the putative suppressant is either an excess of circulating
estradiol or perhaps an excess of trophic hormones.  Estrogens also play a role
in the action of other hormones and trophic factors such as EGF (Kirk!and et
al., 1981; Gardner et al., 1978; Mukku, 1984; Mukku and Stance!, 1985; Hsueh et
al., 1981; Gonzales et al., 1984; Dickson and Lippman, 1987).  These findings
lead to the conclusion that the multiple effects of TCDD could be mediated
through an endocrine mechanism.  The weakness of this assumption is that
-------
 2,3,7,8-TCDD causes effects that appear similar to both  hyperestrogenemia  and
 hypoestrogenemia.   It has been hypothesized that this  apparent paradox is  the
 result of 2,3,7,8-TCDD or the ligand complex preventing  the endogenous
 substrate from interacting "correctly"  with both the active site and  a
 secondary binding  site (Umbreit and Gallo,  1988; Umbreit et al *,  1988).
      Pleiotropism  is not an uncommon finding with molecules such  as hormones or
 in  this  case 2,3,7,8-TCDD.   One has only to review the early experiments on
 multistage mouse skin carcinogenesis of 2,3,7,8-TCDD to  see that  in some cases
 it  inhibited tumor formation by PAH initiators  (DiGiovanni  et al., 1977; Berry
 et  al.,  1978).   It must  be  emphasized that  the  responses in multistage models
 are dependent on time, sequence of  administration,  dose,  and species.  Hence,
 inhibition under some conditions might  have been predictable.  This is
 juxtaposed to the  two-stage liver model  (Pitot  et al., 1980)  in which  it has
 been  shown that orally administered  2,3,7,8-TCDD enhances the tumorigenic
 action of  DEN.  However,  in subsequent  experiments  at  lower doses of 2,3,7,8-
 TCDD, a  parabolic  dose-response curve has been  reported  in  the DEN/2,3,7,8-
 TCDD  initiation-promotion protocol  (Pitot et  al.,  1987).    This paradoxical
 effect is  not well  understood,  but  it does  not  appear  to  be solely the function
 of enhanced  metabolism,  or  Ah  receptor  binding  (Mason  et  al., 1987).   Perhaps
 it is the  result of alteration  of EGF receptors  at  low doses  (Madhukar et al.,
 1984) which  displays  a commonality with  several  steroid  hormone receptors.
     The importance  of these findings to the  approximation  of human and animal
 health risks  from  exposure  to  PCDDs  and  related  molecules cannot be overstated.
 Mathematical  modeling of physiological  phenomena, especially  those related to
 receptor function,  is conducted using the Michaelis-Menten  equation (1913)  as
modified by  Clark  for the "classical" receptor model (1933).  The weight of

                                      10
-------
 evidence for the most prevalent 2,3,7,8-TCDD effects falls into the category of
 the receptor model  (Poland and Knutson, 1982). The recent finding that the
 hepatocarcinogenesis is related to estrogen levels or to the presence of
 functional ovaries  (Goldstein et al., 1987), and that DEN hepatocarcinogenesis,
 in partially hepatectomized rats, is first inhibited then promoted by 2,3,7,8-
 TCDD (Pitot et al., 1980, 1987) indicate that 2,3,7,8-TCDD is not causing its
 myriad of effects in liver by a simple one-step event such as binding to the Ah
 receptor and subsequent induction of cytochrome Pj-450.  However, operationally
 2,3,7,8-TCDD is a potent hepatocarcinogen in some species and strains of
 rodents.                                                          ,
      Risk modeling for carcinogenic xenobiotics has recently been segregated
 into three  classes or types  of models:   physiologically  based pharmacokinetic
 (PBPK)  models  in  which  the body is  considered  to  be a small  group of
 physiological  compartments (Hoel  et  al.,  1983;  Krewski et al.,  1986;  Bischoff,
 1987);  biologically  motivated  models of  carcinogenesis (BMMC)  in which the
 carcinogenic process is  considered to occur through  a series  of  linked
 reactions that  result from two  or more molecular  events  followed by  cellular
 amplification by  "promoter" molecules (Moolgavkar,  1986;  Thorslund et  al.,
 1987; Krewski et  al., 1987); and the linearized multistage model  (IMS) of
 Armitage-Doll as modified  by Crump and Howe (1984)  in which it is assumed that
 a sequence of mutational events occur within a single cell leading to  the
 neoplastic change (Armitage,  1985).
     The model   that  appears to accommodate most of the critical  components from
the biological   data  base on 2,3,7,8-TCDD is the BMMC model, which is generally
referred to as   the Moolgavkar-Venzon-Knudson (M-V-K) model (Moolgavkar and
Venzon,  1979; Moolgavkar and  Knudson, 1981).   This model  allows for several  of

                                     11
-------
the concepts of initiation-promotion-progression, along with the growth-
stimulating role of endogenous substrates such as hormones (Moolgavkar, 1986).
Incorporation of some of the factors necessary for the PBPK model can also be
done using the M-V-K model as modified, or, more correctly, expanded by
Thorslund et al. (1987).  These expansions of the M-V-K model give the risk
assessor a powerful tool for looking at cancer risk mechanistically.
     This option is not available with the IMS model as originally proposed.
The use of the IMS model may not be appropriate for the 2,3,7,8-TCDD data set
since this model assumes an initiating event, such as a point mutation, to
start the process.  However, the IMS model can be accommodated if one
hypothesizes that the initiating event:   (1) is the result of an indirect
action of 2,3,7,8-TCDD through modification of exogenous or endogenous
compounds, (2) that a population of initiated cells exists, or (3) 2,3,7,8-
TCDD leads to focal necrosis which serves as a mitogenic stimuli.
     Recent reports have shown that 2,3,7,8-TCDD and other promoters in liver
enhance stimulation of DNA synthesis in situ, and stimulate repair of
0-6-methylguanine in liver DNA (Busser and Lutz, 1987; Den Engelse et al.,
1986).  Lutz et al. (1984) presented a scheme for promoter potency based on
stimulation of DNA synthesis and the assumption that cell division is a
prerequisite for several stages in the carcinogenesis process.  These reports
indicate that 2,3,7,8-TCDD can act as a complete-indirect-carcinogen, including
promoter activity, despite the lack of DNA binding or direct mutagenesis.  The
sum of all these findings, along with the myriad of toxic responses, suggests a
model for 2,3,7,8-TCDD carcinogenesis in rodent liver as shown in Figure 1.
This model can account for the dose-response data in the bioassays and the
multistage promotion experiments, as well as allow for incorporation into

                                      12
-------
existing risk models, and the scheme is not incongruent with the reports of
decreased tumor formation in some tissues.  If pathway (A) can be verified by
demonstration of reactive intermediates after 2,3,7,8-TCDD treatment, then the
LMS model, slightly modified, can be used.  The preponderance of evidence at
the moment supports a mechanistic model(s) which is at variance with the LMS
model.  However, Figure 1 presents possibilities that are not mutually
exclusive for the existing models.  The scheme also presents several testable
hypotheses which should be examined.
                                     13
-------
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-------
                                   REFERENCES
 Armitage,  P.   (1985)   Environ.  Health  Perspect.  63:195-201.

 Barsotti,  D.A.;  Abrahamson,  L.J.;  Allen,  J.R.   (1979)   Bull.  Environ.  Contam.
      Toxicol.  21:463-469.

 Becker,  P.B.;  Gloss,  B.; Schmid, W.; Strahle,  U.;  Schultz, G.   (1986)   Nature
      (London)  324:686-688.

 Berry, D.L.; DiGiovanni, J.; Juchau, M.R.;  Bracken, W.M.; Gleason,  G.L.;
      Slaga, T.J.   (1978)  Res.  Commun. Chem. Pathol. Pharmacol.  20:101-107.

 Bischoff,  K.B.   (1987)   In:  Drinking water  and health 8:36-64.   Washington,
      DC: National Academy Press.

 Busser,  M-T.;  Lutz, W.K.  (1987)   Carcinogenesis 8:1433-1437.

 Clark, A.J.  (1933)   The mode of action of  drugs on cells.  Baltimore,  MD:
      Williams  and Wilkins.

 Crump, K.S.; Howe, R..B.  (1984)  Risk Analysis 4:163-176.

 Den Engelse, L.; Floot, B.G.J.; Menkveld, G.J.; Tates, A.D.   (1986)
      Carcinogenesis 7:1941-1947.

 Denison, M.S.; Vella,  L.M.;  Okey,  A.B.  (1986)  J. Biol. Chem. 261:3987-3995.

 Dickson, R.B.; Lippman, M.E.  (1987)  Endocr. Rev. 8:29-43.

 DiGiovanni, J.;  Viaje, A.; Berry,  D.L.; Slaga, T.J.; Juchau, M.R.   (1977)
      Bull. Environ. Contam.  Toxicol. 18:552-557.

 Gallo, M.A.; Hesse, E.J.; Macdonald, G.J.;  Umbreit, T.H.  (1986)  Toxicol.
      Lett. 32:123-132.

 Gardner, R.M.; Kirkland, J.L.;  Ireland, J.S.; Stance!, G.M. (1978)
      Endocrinology 103:1164-1172.

 Gierthy, J.F.; Dickerman, H.W.; Seeger, J.L.; Lincoln, D.W.; Martirez,  H.;
      Kumar, S.A.  (1987)  Toxicologist 7:159.

 Goldstein, J.A.; Graham, M.J.; Sloop, T.; Maronpot, R.; Goodrow, T.; Lucier,
      G.W.  (1987)  Dioxin 87, Abstract # MA-07.

 Gonzales, F.;  Lakshman, S.;  Hoath,  S.;  Fisher,  D.A.  (1984)   Acta Endocr.
      105:425-428.

Grossman, C.J.    (1984)  Endocr.  Rev. 5:435-455.


                                     15
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