EPA-453/R-94-085
EXPOSURE AND EFFECTS
OF
AIRBORNE CONTAMINATION
for the
Great Waters Program Report
United States Environmental Protection Agency
22 December 1992
Project Team:
Wayland Swain (Team Lead)
Theo Colborn
Carpi: Bason
Robert Howarth
Lorraine Lamey
Brent Palmer
Deborah Swackhamer
-------
DISCLAIMER
This document was prepared by researchers in Great Waters-
related scientific disciplines, and a draft of this report was
reviewed by an expanded group of scientists at a workshop held in
November 1992 in Chapel Hill, North Carolina. Other workshop
participants included representatives from the U.S. Environmental
Protection Agency, the National Oceanic and Atmospheric
Administration, the International Joint Commission, and the
affected States.
This report has been reviewed by the Office of Air Quality
Planning and Standards, Pollutant Assessment Branch, U.S.
Environmental Protection Agency, and has been approved for
distribution as received from the team of authors. Approval does
not signify that the contents reflect the views and policies of
the U.S. Environmental Protection Agency, nor does mention of
trade names or commercial products constitute endorsement or
recommendation for use.
-------
22 EXPOSURE AND EFFECTS OF AIRBORNE
CONTAMINANTS: PUBLIC HEALTH AND
ENVIRONMENTAL IMPACTS
22.1 Introduction
The chemical properties and the extensive historic utilization of a number of residue-
fonning xenobiotic substances of anthropogenic origin have led to the ubiquitous distribution of
these materials throughout the global environment. This contention is supported by a substantial
body of literature which has documented the presence of anthropogenic contaminants in areas
presumably remote from the direct industrial and/or cultural influences attributable to humans.
These remote sites have included snow in the Antarctic (Peterle 1969; Peel 1975), mammals of
the Arctic (Bowes and Jonkel 1975; Clausen el aL 1973), in the surface waters and atmosphere
above the Sargasso Sea (Bidleman and Olney 1974), rainfall in the South Pacific Ocean
(Benvenue el aL 1972), remote island sites in the North American Great Lakes (Murphy and
Rzeozutko 1977; Swain 1978; Swackhamer si aL 1988; Swackhamer and Kites 1988), and in the
surface waters of most of the world's oceans, including the Atlantic Gulf stream, the Sargasso
Sea, the continental shelves of Iceland, Ireland, Norway, Portugal (Ballschmiter el aL 1981), the
Caspian Sea, the North Pacific and Antarctic Oceans (Zell and Ballschmiter 1980a), the North
Sea and North Atlantic Ocean (Ballschmiter el aL 1978; Zell and Ballschmiter 1980b). Given
this world-wide distribution, it is not surprising, then, that one of the chief mechanisms involved
in the movement of these compounds is atmospheric transport.
Large aquatic and marine ecosystems are morphometrically — and hence, physically,
chemically, and biologically — predisposed to excessive susceptibility to toxic chemical insult.
Many large aquatic, estuarian, or coastal marine ecosystems are geographically located in
physical proximity to large population centers, and hence, pollution sources. Atmospherically-
derived contamination to these systems is quantitatively significant because of the vast surface
areas of these water bodies. Atmospheric inputs are particularly significant to large aquatic and
marine ecosystems, since the contribution is direct, and not filtered through soils and sediments,
as is often the case for tributary derived pollutants (Sonzogni end Swain 1980).
These large ecosystems are frequently oligotrophic in nature, i.e., they are relatively
unproductive with a relatively low autochthonous production of particulate matter. Further,
because of their low suspended sediment load per unit volume, the opportunity for sorption,
scavenging, and subsequent removal to the sediments is markedly decreased. Low solids burdens
and decreased volumetric inputs of particulate matter also diminish the capacity of these systems
to dilute the concentrations of toxic materials once they have been deposited in the bottom
sediments.
Because of the enormous depth of some of these ecosystems, the length of time required
for even the low quantities of particulate matter available to settle to the bottom is excessive,
allowing a considerable period for exposure of fish and other biota to the particulate-borne
contaminants. The increased time of retention of toxic substances in the water column is also
-------
aided by wind-driven circulation, resuspension and mixing in the water column (Sonzogni and
Swain 1980). Hydraulic detention times of the order of decades to centuries have been calculated
for some of these large ecosystems (Quinn 1992). All of these factors tend to increase the
opportunity for exposure of the biota to toxic chemical insult, primarily because natural removal
mechanisms function at such a slow rate.
Finally because of their trophic status, these systems are likely to contain highly sensitive
biota in which one or more life stages may be particularly sensitive to the influence of toxic
contaminants. The question of human exposure potential is also involved, because the biota of
the upper trophic levels are regarded as highly desirable by commercial fish harvesters and sports
and subsistence anglers.
The purpose of this chapter is to examine the existing scientific literature related to
atmospherically transported contaminants and summarize present knowledge about the types and
kinds of chemical contaminants of concern, the pathways and processes involved in exposure,
and the multiplicity of effects associated with these substances. Then, having examined the
present base of information, efforts will be made to identify knowledge gaps and information
deficits. From this basis, future information needs can be identified which will serve to indicate
new or expanded research directions required for the coming decade.
22.2 Elements of Atmospheric Transport
The ubiquitous global distribution of many of the contaminants of concern, particularly
the residue-forming organochlorine compounds, has been well documented. A number of the
compounds commonly included in this group of contaminants of concern have had their North
American production and usage severely curtailed or eliminated in the 1970 to 1983 tune period.
Despite this fact, these compounds continue to be reported in biologic tissues taken from large
aquatic and marine systems, both in North America and throughput the world (Veith el aL 1977;
Norstrom el aL 1980; Schmitt el aL 1981; Schmitt el aL 1985; Ahlborg el aL 1992). The
environmental persistence of these compounds (Ballschmiter el aL 1978) is only a partial
explanation for these continued observations. It is reasonable to expect observations of these
compounds whose biological half-lives are of the order of years to decades to persist in
biological tissue, particularly in long-lived species. However, it is less reasonable to anticipate
that these compounds might be so uniformly observed in fresh mobile sediments and in the water
column itself (Glooschenko el aL 1976; Frank el aL 1977; Swain 1978; Eisenreich and Johnson
1983). Atmospheric transport of residue-forming xenobiotic compounds provides an explanation
for the continued observation of these compounds in a variety of environmental media (Strachan
and Huneault 1979; Eisenreich el aL 1981). The short- and intermediate-range aerial transport
of these substances is well recognized (Olie el aL 1977; Olie el aL 1983; Hutzinger el aL 1985;
Kuehl el aL 1985). Long-range atmospheric movement of the order of hundreds to thousands
of kilometers has frequently been implicated by existing data (Risebrough el aL 1968; Seba and
Prospero 1971, 1972; Spencer 1974; Peakall 1976; Hoff el aL 1992a, b), but only in a few
instances has it been possible to directly associate the observation of the compounds of concern
-------
in atmospheric or precipitation samples with an environmental application or incident (Cohen and
Pinkerton 1966; Rice and Evans 1984; Swain el aL 1986).
Except immediately downwind from a substantial source of contamination, the atmosphere
does not represent a significant reservoir for most organic compounds. To illustrate this facet,
the contemporary burdens of polychlorinated biphenyls (PCBs) for a variety of environmental
media hi the Great Lakes basin are presented in Table 1.
While the atmosphere is not typically a substantial reservoir for contaminants, atmospheric
transport is frequently the major pathway by which contaminants enter marine and large aquatic
ecosystems. The data from the International Joint Commission (1987) suggest the magnitude of
the atmospheric loading of PCBs to the Great Lakes (Figure 1). More than half of the total PCB
loading to the Upper Great Lakes (Lake Superior, 90 percent, Lake Huron, 78 percent, Lake
Michigan, 58 percent) is the result of the direct or indirect contribution of the atmosphere.
Once a compound of concern has entered the atmosphere, either in the form of a
particulate or vapor phase emission, it is possible for these materials to travel great distances.
The transport of contaminants is dependent upon a number of factors including air currents,
particle size, vapor pressure, vapor partitioning, scavenging of particles by water droplets,
washout phenomena, and particle settling (Strachan and Huneault 1979; Eisenreich el aL 1981;
Eisenreich and Johnson 1983; Murphy 1984). While a complete discussion of these factors is
beyond the scope of this review, a number of the major processes are summarized below.
2.2.2.1 Physical Properties and Atmospheric Distribution
The organic compounds of concern have varying physical properties, both by individual
substance and by compound class. However, despite their individual variation, their general
similarities to each other are greater than their differences (Murphy 1984). These organic
compounds tend to form persistent residues in various environmental compartments, including
biota; they tend to have low vapor pressures (< 1Q~5 atm); and they generally have high
solubilities in non-polar liquids and low solubilities in water (< 1 mg/1).
In the atmosphere, trace organic compounds are distributed between the vapor phase and
the particulate, or aerosol, phase. Vapor-aerosol partitioning in the atmosphere is a function of
the individual compound's vapor pressure, the size, type, and surface area of suspended
atmospheric particulates, and the organic content of the aerosol phase. Volatile organic materials,
existing as vapor in the atmosphere, can be either adsorbed on the surface of particles, or
absorbed by non-polar particulates. The quantity of organic compound adsorbed is a function
of the surface area and chemical constituents of the particles in the atmospheric aerosol. The
quantity of organic compound absorbed by non-polar particulate matter is determined by the
quantity of the particulate matter present and the capacity of those particles for absorption, i.e.,
their fugacity (Murphy 1984).
-------
TABLE 1
CONTEMPORARY PCB CONCENTRATIONS IN ENVIRONMENTAL MEDIA
IN THE GREAT LAKES BASIN
LAKE
SUPERIOR
Mean
Range
LAKE
MICHIGAN
Mean
Range
Air
(ng/m3)1
0.2
0.2-0.4
0.3
0.1-1.5
Rain
(ng/L)'
1.0
1.0-5.0
2.0
LAKE
HURON
Mean
Range
0.2
0.15-0.25
LAKE
ERIE
Mean
Range
LAKE
ONTARIO
Mean
Range
0.4
0.4-0.5
2.0
2.0
0.4
0.3-0.5
3.0
Water
(ng/L)
0.2"
0.1-0.3
0.63C
o.3-i.r
0.49e
0.28-0.5T
Sediments
(ng/g)
96
4-12b
81d
l-201d
Sources: a. Eisenreich and Strachan (1992) d. Swackhamer and Armstrong (1988)
b. Eisenreich and Jeremiason (1992) e. Swain el al. (1986)
c. Swackhamer el aL (1992)
-------
FIGURE 1
ATMOSPHERIC LOADING OF PCBs TO THE GREAT LAKES
TOTAL INPUTS, kg/yr
(all sources)
Superior 606
Michigan 685
Huron 636
Erie 2,520
Ontario 2,540
Superior 90%
Ontario
Erie
Small arrows indicate atmospheric contribution falling directly on each lake.
Large arrows denote indirect atmospheric contribution passed down from lakes
"upstream." a = total atmospheric contribution, both direct and indirect.
Source: International Joint Commission (1987)
-------
2.2.2.2 Atmospheric Deposition Processes
There are three dominant processes that transfer organic contaminants from the
atmosphere to marine and large aquatic ecosystems. These are: (1) vapor partitioning across the
air-water interface; (2) dry deposition; and (3) precipitation (wet) deposition. A brief discussion
of the importance of each of these processes is presented below.
Contaminants in the vapor-phase tend to partition directly across the air-water interface.
The tendency to move from one medium to another is based upon the fugacity of the individual
compound (Mackay 1979; Mackay and Patterson 1981). The fugacity of a compound is a
measurement of its tendency to escape from a particular medium into another physical phase or
medium. In short, the fugacity of a material is its tendency to partition from one medium to
another. If a vapor-phase, airborne contaminant immediately above the surface of the water is
at equilibrium with both phases, the air and the water, the fugacity of that contaminant is the
same, and no vapor-phase partitioning will occur (Murphy 1984). However, if the fugacity of
one phase exceeds that of the other, the contaminant of concern will tend to partition from the
phase with the higher fugacity toward the lower.
An example of vapor-phase transfer has been provided by Eisenreich and Looney (1982).
These authors made very careful measurements of PCBs in the water column and in the
atmosphere under stable atmospheric conditions over Lake Superior. They found significantly
higher concentrations of PCBs in the water layer at the surface than in those layers deeper in the
water column. These authors reported that atmospheric vapor inputs of low-volatility PCBs were
responsible for maintaining the gradient observed in Lake Superior.
The settling of particles onto a surface hi the absence of a precipitation event is referred
to as dry deposition. Slinn el aL (1978) and Slum and Slum (1980) have considered dry
deposition to bodies of water. These authors note that the deposition velocity or rate of
deposition of an organic compound is a function of the size particle to which it is sorbed. The
smallest of the particles have aerodynamic diameters, known as mass median diameters (mmd),
of < 0.3 um. While more dense than air, these particles are small enough to be moved about by
Brownian diffusion. Since these particles are unaffected by a gravitational component, their
deposition is independent of the orientation of the surface with which they collide. The next
larger size particles are those with aerodynamic diameters in the range of 0.5 to 2-5 um. These
particles are deposited on surfaces by impaction. Particles with mmd values greater than 2-5 um
are too large to be seriously influenced by air molecules and Brownian movement. Because their
mass is greater, gravity imparts a net downward movement on these particles known as a
deposition velocity. Gravitational sedimentation from the atmosphere is the principal removal
mechanism for these particles. Ultragiant particles, those particles with aerodynamic diameters
(mmd) greater than 10 um, also have an increased deposition velocity as a function of their
increased mass. The relationships of particle size and deposition velocities are shown in Table
2.
Andren and Strand (1981) have shown that 70 percent of the total organic carbon
associated with airborne particulate matter over Lake Michigan is transported by particles < 1.0
-------
um in size. Because of the greater surface-to-volume ratio and higher organic content of
particles in this size range, Doskey and Andren (1981a, b) reported that polychlorinated biphenyls
(PCBs) are associated with these submicron sized particulates.
Precipitation in the form of rain and snowfall is another major mechanism for the
deposition of organic contaminants to large water bodies. In the atmosphere, aerosol particulates
are concentrated and removed by a variety of events related to precipitation. Atmospheric
particulates serve as droplet condensation nuclei forming clouds. Cloud droplets formed in this
manner may also scavenge additional particulate matter from the air mass. Scott (1981) reports
that the coalescence of approximately 106 cloud droplets in a liter of air can result in an increase
in concentration of trace organic compounds by 10* to 106 in the resulting precipitation by this
mechanism. Further, if an organic compound has a tendency to partition into water, vapor phase
compounds in the atmosphere can be substantially higher in precipitation.
TABLE 2
RELATIONSHIP OF PARTICLE SIZE TO
DEPOSITION MECHANISM AND DEPOSITION VELOCITY (¥„)
PARTICULATE
MASS MEDIAN
DIAMETER
(/tin)
<0.3
0.5 to 2 - 5
>2-5
DEPOSITION
MECHANISM
Brownian diffusion
Inertial Impaction
Gravitational
Sedimentation
APPROXIMATE
DEPOSITION
VELOCITY
(m/s)
Isotrophic
(- 0.005)
< 0.002
> 0.005
Sources: Eisenreich el aL (1981) and Murphy (1984)
-------
The results of field studies suggest that the bulk of the trace organic contaminants in
precipitation is associated with particulate matter. Hence, the majority of the contaminant
transferred to a large water body will be deposited in the early stages of a precipitation event.
The first few millimeters of precipitation contain relatively high concentrations of the
contaminant as a result of atmospheric washout, while the remainder of the precipitation event,
containing much reduced concentrations of the contaminant, serves essentially as dilution for the
earlier deposition (Strachan and Huneault 1979; Murphy and Rzeszutko 1977).
The amount of variation in contaminant levels in individual precipitation events has been
demonstrated by Murphy (1984) and Swain el aL (1986). The variation in precipitation inputs
of PCBs to the Great Lakes has been summarized in Table 3.
Swackhamer and Armstrong (1986) have demonstrated the relative importance of these
major removal processes by creating a mass balance for PCBs in Lake Michigan. These authors
have demonstrated that, for PCBs, the following mass removal hierarchy exists:
wet washout (particles) > wet washout (vapor) > dry deposition (particles).
2.2.2.3 Atmospheric Deposition
Having reviewed the literature for the preceding decade, Eisenreich el aL (1981)
summarized the trace organic contaminant concentrations in the atmosphere and in precipitation
in the Great Lakes basin. Their findings are presented hi Table 4. From the mean values
reported (Table 4) for contaminants in air and precipitation, the equations for wet and dry flux
were used to achieve an estimate of annual atmospheric loadings to the Great Lakes for the time
period. The Eisenreich el aL (1981) data for total annual atmospheric loadings for a variety of
atmospherically-borne pollutants are presented in Table 5.
2.2.2.4 Relationships to Water Quality Criteria
Over the last two decades, the United States Environmental Protection Agency (USEPA)
has developed water quality criteria for nearly 200 chemical entities and substances. The specific
value for each substance adopted by USEPA was based upon exhaustive examination of the
scientific literature and knowledge of that particular chemical entity. From that knowledge,
criteria were developed designed to be protective under specific scenarios, e.g., acute or chronic
criteria for freshwater ecosystems as contrasted to the acute or chronic values for marine systems.
In addition, human health criteria were established based upon a lifetime one in a million risk
of cancer. The water quality criteria values for a number of contaminating compounds of
concern in the world's great waters are presented hi Table 6.
Subsequent to the earlier Eisenreich el aL (1981) study of atmospherically transported
contaminants (Table 6) (Eisenreich and Strachan 1992) estimated that transport and deposition
8
-------
TABLE 3
VARIATION IN PRECIPITATION INPUTS OF PCBs TO THE GREAT LAKES
LOCATION
Picton (L. Ontario)
Point Pelee (L. Erie)
Goderich (L. Huron)
Nipigon; Batchawana Bay
(L. Superior)
Chicago (L. Michigan)
Chicago (L. Michigan)
Waukegan (L. Michigan)
Point Betsie (L. Michigan)
Whitestone Point (L. Huron)
Tawas Point (L. Huron)
Lake Superior
Lake Ontario
Saginaw Bay (L. Huron)
Duluth (L. Superior)
Isle Royale (L. Superior)
PCB
CONCENTRATION
(ng/1)
32
9
11
26
104
75
46
12
13
18
38
43
25
50
230
VOLUME
OF
PRECIPITATION
(cm)
16
6
11
10
39
20
55
63
34
—
—
—
—
13
25
METHOD
Event
Event
Event
Event
Event
Event
Event
Event
Event
Snow Cores
Snow Cores
Snow Cores
Ice Cores
Snow Event
Snow Event
REFERENCE I
Strachan and Huneault (1979)
Strachan and Huneault (1979)
Strachan and Huneault (1979)
Strachan and Huneault (1979)
Murphy and Rzeszutko (1977)
Murphy el aL (1982)
Murphy el aL (1982)
Murphy et aL (1982)
Murphy el aL (1982)
Strachan and Huneault (1979)
Strachan and Huneault (1979)
Strachan and Huneault (1979)
Murphy and Schinsky (1982)
Swain (1978) I
Swain (1978) |
Source: Murphy (1984)
-------
TABLE 4
AIRBORNE TRACE ORGANIC CONCENTRATIONS IN
THE GREAT LAKES ECOSYSTEM
Air Precipitation
Total PCB
Total DDT
a-BHC
Y-BHC
Dieldrin
HCB
p,p'methoxychlor
a-Endosulfan
6-Endosulfan
Total PAH
Anthracene
Phenanthrene
Pyrene
Benz[a]anthracene
Perylene
Benzo[a] pyrene
TOC
DBP
DEHP
Range Mean Range Mean
(ng/mj) (ng/L)
0.4-3
.01-.05
.25-0.4
1-4
.01-0.1
.01-0.1
—
—
—
10-30
0.1-1
0.1-1
0.1-4
0.1-1
0.1-2
0.1-2
2-15 x 103
0.5 -5
0.5-5
1.0
0.03
0.3
2
0.05
0.05
1
1
1
20
0.6
0.6
1.1
0.5
.06
1
9X103
2
2
10-100
1-10
1-35
1-15
0.5-30
0.5-30
1-20
1-10
1-12
50-300
1.3-2.3
2.0-2.3
1.3-4.5
2.6-3.1
—
0.1-3.1
1-5 x 105
4-10
4-10
30
5
15
5
2
2
8
2
3
100
2
2
2
.3
1
2
2x 106
6
6
Source: Eisenreich el aL (1981)
-------
TABLES
TOTAL DEPOSITION OF AIRBORNE TRACE ORGANIC
COMPOUNDS TO THE GREAT LAKES (metric tons per year)
COMPOUND
Total PCB
Total DDT
a-BHC
Y-BHC
Dieldrin
HCB
p,p'methoxychlor
a-Endosulfan
B-Endosulfan
Total PAH
Anthracene
Phenanthrene
Pyrene
Benz[a]
anthracene
Perylene
BeHZo[a]pyrene
DBP
DEHP
Total Organic
Carbons
LAKE
Superior
9.8
0.58
3.3
15.9
0.54
1.7
8.3
7.9
8.0
163
4.8
4.8
8.3
4.1
4.8
7.9
16
16
2X205
Michigan
6.9
0.40
2.3
11.2
0.38
1.2
5.9
5.6
5.6
114
3.4
3.4
5.9
2.9
3.3
5.6
11
11
1.4 x 10s
Huron
7.2
0.43
2.4
11.6
0.55
1.2
6.1
5.8
5.8
118
3.5
3.5
6.1
3.0
3.4
5.8
12
12
1.5 xlO5
Erie
3.1
0.19
1.1
5.0
0.17
0.53
2.6
2.5
2.5
51
1.5
1.5
2.6
1.5
1.5
2.5
5.0
5.0
.66X105
Ontario
2.3
0.14
0.77
3.7
0.13
0.39
1.9
1.8
1.9
38
1.1
1.1
1.9
1.1
1.1
1.8
3.7
3.7
.46X105
Source: Eisenreich si aL (1981)
-------
of a number of toxic substances to the Great Lakes region. Appendix III of their report contains
a summary of the recent measurements of contaminant concentrations in rainfall. These data are
also presented in Table 6 for comparison with the USEPA Water Quality Criteria for surface
waters.
In comparing the concentrations of contaminants in rainfall with the water quality criteria
for surface waters, it must be recalled that wet deposition is only a fraction of the total
contribution of the atmosphere to the world's great waters. Dry deposition is also responsible for
addition of substantial quantities of some contaminants. Calculation of the total flux to
waterbodies for each of these compounds is beyond the scope of this paper. However, the
averages of measured concentration in precipitation are sufficient to suggest the magnitude of the
problem of atmospherically transported contaminants.
The data in Table 6 suggests that in four other instances, the concentrations of
contaminants hi rainfall exceeded the human health criteria for 10"6 cancer risk. The compounds
hi this group consisted of polychlorinated biphenyls (PCBs), dieldrin, dioxin, and DDT. The
mean precipitation values of two additional substances, hexachlorobenzene and chlordane, are the
same order of magnitude as the published human health criteria. The value of the alpha isomer
of hexachlorocyclohexane (HCH) hi rainfall exceeds the recalculated value for human health
related to water and organisms, as does dieldrin hi precipitation. Four other compounds, DDT,
toxaphene, benzo(a)pyrene, and chlordane either approach the recalculated human health criteria
values in rainfall, or the average rainfall values are of the same order of magnitude as the human
health 10"6 cancer risk values recalculated from the IRIS database.
It is clear that the water quality criteria are intended to be applied to the world's great
waters and to other bodies of surface water. It is alarming to discover that the precipitation
which drives these bodies of water, directly or indirectly, contains average concentrations which
exceed or approach criteria one or more water quality criteria values. In fact, of the organic
compounds examined, only the gamma isomer of HCH meet all the concentration requirements.
Three additional substances also meeting all the criteria limits were metals, i.e., arsenic,
cadmium, and lead. If nothing more, this comparison is .an indication of the extent of the
problem posed by atmospherically transported substances.
22.3 Compounds of Concern
2.2.3.1 Identification of Compounds of Concern
There are over 65,000 chemicals registered for current use hi the United States, with new
ones added continuously. Many of these chemicals are released into the environment by
discharges into air, water, land, sewer systems, or subsurface. More than 1000 chemicals have
been identified hi the waters of the Great Lakes. The Toxic Release Inventory, established as
part of the Emergency Planning and Community Right-To-Know Act, requires industry that
report on over 300 chemicals and chemical categories. Air emissions of these chemicals account
for more than 40 percent of all emissions to all media (EPA 1991). In an attempt to reduce these
emissions, the Clean Air Act Amendments of 1991 identify 189 hazardous air pollutants for
regulation by the EPA.
12
-------
TABLE 6
CONCENTRATIONS OF THE COMPOUNDS OF CONCERN IN PRECIPITATION
COMPARED WITH THE USEPA WATER QUALITY CRITERIA
(All Values /
-------
TABLE 6 (Cont.)
Compound
Y HCH (BHQ
Chlordane
Lead
Cadmium
Mercury
Arsenic
Copper
Zinc
N, P
Acute
Criteria
(FRESH)
2.0
2.4
83.0+
3.9+
2.4
—
Chronic
Criteria
(FRESH)
0.08
0.0043
3.2+
1.1+
0.012
—
Acute
Criteria
(MARINE)
0.16
O.OSl
220
43.0
2.1
—
Chronic
Criteria
(MARINE)
—
0.004
8.5
9.3
0.025
—
Human Health W* Risk Level for Carcinogens
Published Criteria
WATER AND
ORGANISMS
0.0186
0.00046
50.0
10.0
0.144
0.0022
ORGANISMS
ONLY
0.0625 .
0.00048
—
—
0.146
0.0175
Recalculated Values
Using IRIS database
WATER AND
ORGANISMS
0.019
0.00058
—
10.0
0.14
0.018
ORGANISMS
ONLY
0.063
0.00059
—
170
0.15
0.14
Estimated
Mean and
(RANGE) of
Concentra-
tions in
Rainfall
.0034
(0.001-0.01)
0.0002
(0.00003-
0.00045)
.004
(0.0007-
0.012)
0.002
(ND-0.007)
0.025
(0.003-
0.213)
0.0003
(0.0001-
0.0004)
Sources: USEPA 1991 Water Quality Criteria from Health and Ecological Criteria Division of the Office of Science and Technology (G. Glass, personal communication).
Rainfall Concentrations from Eisenreich and Strachan 1992.
(P)
Proposed criterion
Insufficient data to develop criterion, value presented is the L.O.E.L. (lowest observed effect level)
Comparative rainfall data are for all tetrachlorodibenzo(p)dioxins
Hardness dependent criteria (100 ug/l CaCo, used)
-------
Not all of these chemicals present equal degrees of hazard to the environment, as they
have differing chemical behaviors, fates, exposure concentrations, and toxic effects. Thus this
formidable list of contaminants can be characterized by level of concern based on the above
differences. The characteristics that give rise to greater concern include persistence in the
environment, measurable toxicity, and the potential for chemicals to build up in animal tissue
such that the concentrations increase within the food web. The same chemical properties that
cause persistence often contribute to toxicity, to long range transport, and to lipophilicity which
allows them to bioaccumulate. This section focuses on those persistent toxic chemicals that move
from air to water and can accumulate in food webs.
Many chemicals in the atmosphere have short lifetimes, due to transformation processes
such as photolysis or reactions with radicals, or due to rapid removal processes that deposit the
contaminant close to its source. Examples of these would include benzene (former) and lead
(latter). Chemicals may have high Henry's Law constants such that they do not readily partition
from air to water (for example, toluene). The Great Lakes Water Quality Board of the
International Joint Commission (GLWQB, 1987) prioritized the contaminants ("IJC Critical
Pollutants") in the Great Lakes according to persistence, lipophilicity, and toxicity. These
chemicals are of concern in all the Great Waters of the U.S., and are not specific to the Great
Lakes. The chemical properties that control the behaviors of persistence, lipophilicity, and
toxicity are vapor pressure, aqueous solubility, and the octanol-water partition coefficient, Kow.
Compounds with low Henry's Law constants (approximated by the ratio of vapor pressure to
aqueous solubility) readily partition from the gas phase to water, and do not readily revolatilize
(Mackay 1982). Compounds with low solubilities are usually associated with particles once they
are in water, and thus may not be available to undergo transformation reactions. Compounds
with high Kows are lipophilic and readily accumulate in fat or lipid tissue of plants and animals.
The pollutants considered to be of greatest concern in Great Waters areas are shown in Table 7,
along with their physical properties.
2.2.3.2 Occurrence, Prevalence, and Distribution
The compounds of concern are generally found in the vapor phase or on submicron
atmospheric particles such that they can be carried long distances from their point of origin and
become well-mixed within a given air mass. Furthermore, many of these chemicals no longer
derive from atmospheric point sources, but instead are part of a ubiquitous baseline contamination
of the atmosphere. Examples include PCBs, PCDDs/DFs, DDT and the other organochlorine
pesticides. Many of these are no longer manufactured in this country, but can be found in remote
environments as a result of persistence and long-range atmospheric transport and deposition.
One of the major sources of PCDDs/DFs is waste incinerator emissions, which are found
throughout the country in rough proportion to population. Thus these chemicals have a fairly
constant source which increases near urban areas. There are instances of local or point source
"hot-spots" for some chemicals of concern (e.g., metals concentrations near smelters; metals and
organic chemical concentrations in or adjacent to urban areas); generally the bodies of water
under consideration receive atmospheric loadings representative of the entire region. For
15
-------
instance, PCB concentrations in air are similar in an east-to-west transect over Lake Superior,
which are similar in concentration to measurements over Lake Huron. Concentrations over
southern Lake Michigan are only slightly higher, likely due to the proximity and influence of
Chicago (see Figure 2) (Eisenreich and Strachan 1992). Thus away from large point sources,
concentrations are similar across long distances. Concentrations of PCBs in air over Chesapeake
Bay are also similar to those over the Great Lakes (Baker, unpublished data). In Chesapeake
Bay, chemicals of concern include PCBs, phthalate esters, PAHs, and heavy metals (copper, zinc,
and lead) (Helz and Huggett 1987). Helz and Huggett (1987) and Wright el aL (1992) provide
an extensive review of the field and laboratory studies which describe wildlife health disturbances
observed in Chesapeake Bay and its tributaries. Atmospheric monitoring of these contaminants
by Atmospheric Environment Service of Environment Canada indicates that seasonal variations
in concentrations are often greater than geographical differences (Hoff el aL 1992a, b). Elevated
concentrations due to urban or highly industrialized areas are highly localized. It must be kept
in mind, however, that the atmospheric component of urban area sources to overall loadings of
contaminants to water bodies may be substantial despite the confined geographical area, due to
prevailing wind directions (e.g., the effect of Chicago on southern Lake Michigan). Quantitative
estimates of the urban effects on atmospheric loadings of these contaminants to large water
bodies are lacking. First order estimates of atmospheric loadings and the relative importance of
atmospheric loads compared to non-atmospheric loads have been made for the Great Lakes
(Strachan and Eisenreich 1988; Eisenreich and Strachan 1992).
16
-------
PO
COMPOUND
PCBs
B(a)P
Dieldrin
HCB
TCDD
TCDF
DDT
Toxaphene
a-HCH
g-HCH
Chlordane
nonachlor
Lead
Cadmium
Mercury
Arsenic
LLUTANTS OF <
AND THEIR PHI
VAPOR
PRESSURE
(atm)
6 x 10-" to
1 x 10-5
3 x ID'12 to
7.2 x ID'12
4.1 x ID'9
(20 C°)
1.5 x 10-8 to
3.1 x ID'7
2 x ID'12 to
1.3 x 10-9
2 x ID'" to
1.2 x 10-9
2.5 x lO'10
(20 C°)
2.6 x 10" to
5.3 x lO"
(20 C°)
2.89 x 10-4
(20 C°)
1.24 x 10-8
(20 C°)
2.9 x 10-8 to
3.8 x 10-8
n.a.
TABLE?
:ONCERN IN THE GREAT WATERS
fSICAL CHEMICAL PROPERTIES
SOLUBILITY
(mg/L)
1.8 x lO'5 to 4
0.004 to 0.0063
0.186
0.005 to 0.008
7.2 x 10-6 to
0.002
0.0004 to
0.0035
0.0031 to
0.0034
0.3
0.088
17
0.056
n.a.
KOW
1.99 x 104 to
1.38 x 109
1.1 x 104 to
3.2 x 106
1.2 x 104
1.35 x 104 to
3.16 x 105
2.4 x 105 to
3xl08
1.35 x 104 to
3.16 x 105
1.54 x 106
2X103
2.88 x 103 to
7.08 x 103
1.99 x 103 to
4.07 x 103
3.47 x 10s
n.a.
REFERENCE
Mackay el aL
1992
Mackay el aL
1992
Dynamac Corp.
1989
Mackay el aL
1992
Mackay el aL
1992
Mackay el aL
1992
Verschueren
1983
Clement Assoc.
1990
Clement Assoc.
1989
Clement Assoc.
1989
Clement Assoc.
1989
-------
2.2.3.3 Exposure Routes, Pathways, and Processes
Once chemicals are delivered to water surfaces by atmospheric deposition, they are subject
to a number of additional other physical, chemical, and biological processes before impacting a
biological receptor. A thorough discussion of these processes hi sufficient detail is beyond the
scope of this chapter, however, the reader is referred to a recent review of organic contaminant
behavior hi lakes (Swackhamer and Eisenreich 1991). A brief outline follows. Once in the water
column, contaminants will partition thermodynamically between particles (suspended sediment),
suspended erosional material, phytoplankton, detritus, etc.), dissolved organic material, and it's
truly dissolved form. Hydrophobic organic compounds may be entrained and concentrated at the
air-water interface known as the surface organic microlayer, a region tens to hundreds of microns
thick consisting of high molecular weight macromolecules having both polar and non-polar
functionalities. While contaminant concentrations hi the surface organic microlayer may be
enriched relative to the water column, the mass of contaminant bound up in the microlayer is
small overall.
Most of the organic contaminants and metals of concern have high particle-to-water
partition coefficients (Kp). The fate of the chemical, its persistence in water, and it's availability
to biota are affected by it's distribution between particles, dissolved organic carbon/colloids, and
the dissolved phase. For instance, particle-bound contaminants will deposit to and accumulate
in sediments. The exposure to chemicals by organisms thus is largely controlled by the phase
of the chemical, and it's bioavailability in that phase. The major exposure routes of aquatic
pollutants include exposure directly from the dissolved phase in water and from consuming
contaminated food of aquatic origin. Exposure from DOC or colloid-associated contaminants
is less important (see below). Exposure from water would include dermal exposure by humans,
gill uptake by fish, equilibrium with surrounding water by zooplankton, and sorption to surfaces
of aquatic plants. Exposure by food consumption occurs through both the pelagic and benthic
food webs. Contaminants associated with sediment are grazed by benthic organisms and bottom-
feeding fish; contaminants associated with phytoplankton are grazed by herbivores. These trophic
levels can then be consumed by higher trophic levels, all the way up to wildlife and humans.
Bioaccumulation is the process by which an organism takes up chemical both from water and
from food; bioconcentration describes the uptake of chemical from water only. The ratio of
contaminant concentration in organism to that hi water is known as the bioaccumulation factor.
When the bioaccumulation factor is greater than that predicted by thermodynamic equilibrium
between organism and water (the bioconcentration factor), biomagnification is said to occur.
Bioaccumulation in a pelagic food web is depicted in Figure 3. Thus the type of exposure route,
and the relative importance of each, differs for different receptor organisms.
Phytoplankton accumulate contaminants only from water; fish can accumulate them from
transport across the gill membrane and by assimilation of contaminated food (the food
concentration is dependent on trophic level); human and wildlife exposure is from water
consumption and ingestion of contaminated fish (additional non-aquatic routes of exposure are
also possible, such as inhalation or other food sources). Because of biomagnification of
lipophilic compounds within the food web, top predator exposures hi pelagic food webs are
18
-------
dominated by food consumption rather than from water exposure. For instance, top predators
(lake trout) (Sylvelinus namaycush) in Lake Michigan are estimated to get 99 percent of their
PCS body burden from the food web (Thomann and Connolly 1984). Mackay and coworkers
(Mackay el aL 1985) have modeled TCDD exposure to humans, estimating that the major
exposure route would be from consuming contaminated fish (Figure 4). Note that contaminants
that may be at very low or trace concentrations in water may still be of concern because the
biomagnification that can occur within the food web greatly enhances pollutant exposure.
The actual, "effective", concentration of a contaminant is that fraction of contaminant that
is actually biologically available. Bioavailability is affected by the water-particle partitioning
of the chemical, and by the physical and chemical characteristics of the water body. For toxic
metals, the bioavailable form of the metal is affected by pH, temperature, DO, salinity, redox
conditions, and complexation reactions. Bioavailability of organic compounds is affected by
complexation to DOC (Landrum el aL 1987). There are obvious differences in salinity (and thus
possibly exposure and uptake) between marine and freshwater aquatic systems; salinity gradients
also exist in estuarine systems such as Chesapeake Bay that vary with time and space, as a
function of tides and meteorology. Temperature variations in time, geographical location, and
depth of water column occur across all water bodies of concern, and may affect exposure and
uptake. Likewise, variations in Ph occur on the micro and macro scales in response to physical,
chemical, and biological processes. The effects of these parameters on chemical speciation,
complexation, partitioning, and bioavailability are understood to some extent but will not be
reviewed here. A full review of the bioavailability literature is beyond the scope of this report,
but EPA is encouraged to include such a discussion in future technical support documents.
Temperature, pH, DO and salinity may also alter the internal physiological response of
the organism to the contaminant, although little is known on this subject. Potential effects might
include alterations in cellular transport, membrane permeability, ionic balance, kinetics of the
response, diffusivity of the chemical, and receptor binding. These require much further study.
19
-------
FIGURE 2
ATMOSPHERIC CONCENTRATIONS OF PCB - FALL AND SPRING 1991-92
Concentrations of PCBs in air over open waters of the Great Lakes (ng/m3) from fall and spring, 1991-1992. Data are from
two to four sites per lake, with each site indicated as a bar on the bar graphs. The spatial distribution of sample locations
for each lake, from top to bottom, is as follows: Lake Superior, west to east; Lake Michigan, north to south; Lake Huron,
north to south; Lake St. Glair, top is Lake St. Clair, bottom is Detroit River; Lake Erie, east to west; Lake Ontario, east to
west.
Source: Hombuckle, KL, and Eiscnreich, S.J., Gray Freshwater Biological Institute, University of Minnesota,.unpublished data.
-------
FIGURE 3
FOODCHAIN BIOACCUMULATION
Small Fish
(alewives, chubs,
perch,._)
4 Benthic
I Invertebrates water *r
V AkA^ /
Large Fish
(lake trout, walleye,
Bass,.„)'
Sediments
Humans
Source: Adaptation from WI Sea Grant (1976)
-------
It should be noted that all of the literature reviewed on effects in the field is for northern
temperate climates, and may not be fully representative of the effects in aquatic systems in other
climates, such as southern California estuaries, the Gulf of Mexico, or the coastal estuaries of
Florida. Additional field and experimental work is needed in these areas to document different
physical and chemical environments on the effect of contaminants on organisms.
Uptake by animals is affected by the assimilation efficiency of the compound across the
gut, the respiration rate (for fish), the metabolic rate, and the egestion rate. The physical form
of the contaminant also is important. For instance, the dissolved chemical may be more readily
taken up than the same concentration of chemical associated with particles. A quantitative
understanding of the effects of these parameters on bioavailability is largely lacking. For
instance, the assimilation efficiencies for the vast majority of chemicals for most fish species are
unknown.
An accurate characterization of the effective concentration of contaminant is a critical link
in demonstrating the connection of atmospheric deposition to water, to organism exposure, to
toxic response. Other factors in this linkage will affect the toxic response of an organism. These
include the threshold does required to elicit a response (chemical and organism specific), and the
kinetics of the response.
The linkage of contaminant deposition to effect has been clearly demonstrated for nitrogen
in esruarine systems; it is less clear for the toxic metals and hydrophobia organic compounds.
The litany of effects discussed in the next section are potential effects; the demonstration of
cause-effect is implicated in the Case Studies in Section 2.2.5, and in the field evidence
presented in Section 2.2.4.
The distribution of contaminants between dissolved and particulate phases affects both
bioavailability, and the extent to which contaminants are accumulated in food webs relative to
other fate pathways. In open waters, much of the particulate phase is composed of
phytoplanktoh. In highly productive waters, hydrophobic contaminants associated with
phytoplankton will be removed by sedimentation and buried in the bottom sediments, while less
productive waters, a greater percentage of the phytoplankton will be grazed and the associated
contaminants transferred preferentially to the food web. Thus, phytoplankton can play a key role
in the bioaccumulation process and in affecting exposure of higher organisms to contaminants.
In addition, contaminants can effect phytoplankton primary production and food web
structure. Early studies on the effects of PCBs and DDT on marine phytoplankton show that
species composition of mixed cultures can be altered as sensitive vs. resistant species and small
vs. large species are differently affected. PCB (at 25 ppb) and DDT (at 50 ppb) inhibits growth,
in pure cultures, of the marine diatom Thalassiosirapseudonana, but not the more resistant green
alga Dunaliella tertiolecta (Mosser el aL 1972). When placed in mixed cultures, the sensitivity
of T. pseudonana increased such that its growth was inhibited at PCB concentrations that showed
no effect in pure cultures. This result may be due to limited nutrient availability. That is, when
uninhibited, T. pseudonana assimilates more nutrients than D. tertiolecta because of its greater
22
-------
rate of growth. However, when T. pseudonana is impaired by DDT or PCB, more nutrients are
available to the resistant D. tertiolecta for assimilation. In this way, nutrient availability plays
a key role in determining the effects of chemicals on food web structure. A slow growing, less
abundant, resistant species may become more prominent at the expense of a sensitive species
following chemical exposure. PCBs may impair the growth of T. pseudonana by inhibiting
membrane-bound enzymes involved in nitrogen metabolism (Fisher 1975).
In 1975, Fisher determined that growth, rather than photosynthetic capability, was reduced
in marine algae following PCB (10 ppb) and DDT (50 ppb) exposure. The 72 percent inhibition
of T. pseudonana culture and the 84 percent inhibition of S. costatum culture photosynthesis by
DDT were a result of growth inhibition rather than photosyntheric inhibition. Fisher therefore
concluded that total marine photosynthesis will not show dramatic decline however, the
replacement of sensitive species by dominant species will result in a qualitative rather than
quantitative alteration of herbivores' food supply and, subsequently, the marine food web (Fisher
1975). This alteration could prove dramatic if the sensitive species are a primary food source
for herbivores.
Moore and Harris (1972) also describe a parallel decline in photosynthesis and growth of
natural marine phytoplankton communities following exposure to p,p'-DDT (5 ppb) and 2,4-D
(7 ppb). They also noted that the compounds Aroclor 1242 and Aroclor 1254 were more toxic
to phytoplankton than were the pure compounds, DDT or 2,4-D. Like Mosser el aL (1972), they
noted that organochlorines are more acutely toxic hi mixed cultures than hi single species
cultures.
Harding (1976) noted that phytoplankton photosynthesis may be affected by temporal and
geographical differences due to variations hi salinity, temperature, particulate composition,
nutrient levels and phytoplankton community composition. In the northern Adriatic Sea, PCBs
reduced phytoplankton photosynthesis at 10 ppb; the magnitude of reduction differed with region
and season. In Long Island Sound, two species of Thalassiosira showed inhibited growth and
photosynthesis following a single dose of 10 ug/liter PCB. However, within a few days, the rates
of growth and photosynthesis equalled and surpassed those of the control signifying this species'
ability to completely recover from PCB exposure. Inhibition of photosynthesis is believed to be
due to reduced levels of chlorophyll-a per PCB-treated cell (Powers el aL 1977).
In this experiment, all cell sizes exhibited a reduction to 30 percent of the control
biomass. Because a full recovery of biomass would require several days, in the natural
environment this period of time may suffice for the less dominant, faster-growing and more
resistant species to establish themselves, thereby changing community structure. Also, a period
of days without these essential algae could have a negative impact on herbivore populations.
A study of Long Island Sound natural phytoplankton assemblages also showed a reduction
and recovery of growth after exposure to PCBs at concentrations of 1 or 10 u,g/day (O'Connors
el aL 1978). Rate of recovery increased with higher concentrations. Unlike the above
experiment, effects differed with cell size. Treatment of communities with one u,g/liter PCB
23
-------
affected particles larger than nine um BSD for three days, but smaller particles were unaffected.
Treatment with 10 u.g/liter PCB suppressed small and large particles with a recovery of small
particles within three days. Therefore, large diatoms are more sensitive to PCBs than are smaller
diatoms. PCBs also favored smaller algae in a study of estuarine phytoplankton exposed to five
or 10 jig/liter of Aroclor 1254 (PCB) (Biggs el aL 1978). These results further contribute to the
possibility that organochlorines can affect species composition thereby altering entire oceanic
food webs. Large phytoplankton forming short food chains tend to produce harvestable fish
whereas small phytoplankton believed to produce longer food chains result in "ecosystems
containing numerous ctenophores, jellyfish, and other gelatinous predators" (O'Connors el aL
1978).
Other chemicals which can affect the growth rate and carbon uptake of marine
phytoplankton include chlordane (lOng/liter) (Biggs el aL 1978), Di-n-butyl Phthalate (Acey el
aL 1987) and polynuclear aromatic hydrocarbons (Riznyk el aL 1987).
Effects of contaminants on freshwater algae are similar to marine plankton in that
sensitivity and resistance differ with species. Up until the early 1980's most research was
conducted on marine plankton, with the majority focusing on PCBs. Later research incorporated
insecticide and herbicide effects on stream and lake communities.
The effect of PCBs on freshwater phytoplankton from oligotrophic and eutrophic lakes
appears to be dependent on the density of plankton cells (Sodergren and Gelin 1983). This may
be due to a threshold under which the level of PCBs accumulated per cell do not affect carbon
fixation rates. Therefore, more resistant species are able to assimilate certain PCB concentrations
with only a temporary decline in photosynthetic rate. Phytoplankton in an oligotrophic lake in
Sweden were more sensitive to PCBs (26 ug/liter) than phytoplankton in eutrophic lakes since
oligotrophic phytoplankton did not adapt 16 hours after addition of PCBs (26 ug/liter) than
phytoplankton in eutrophic lakes since oligotrophic phytoplankton did not adapt 16 hours after
addition of PCBs. A 70 percent reduction in carbon fixation rates occurred during the spring and
a 57 percent reduction occurred during the summer (Sodergren and Gelin 1983). Further
reduction was noted after 16 hours.
In contrast, eutrophic lake phytoplankton, following a large spring bloom of the diatom
Stephanodiscus hantzshii, suffered a 15 percent reduction in primary productivity following PCB
addition. Photosynthesis rates showed greater reduction during the autumn when phytoplankton
biomass was smaller. Of the total amount of the 26 ug/liter PCBs added to the eutrophic lake
phytoplankton samples, 46 percent was found in the algae during the spring and 30 percent in
the autumn (Sodergren and Gelin 1983).
Transmission electron microscopy studies of algae ultrastructure following PCB exposure
showed that the chloroplast is the organelle most sensitive to PCBs. Chlorella fusca var.
vacuolata, Scenedesmus quadricauda, and Scenedesmus obliquus all showed disruption of the
chloroplast after a 48 hour exposure to one ug/ml of PCB (Mahanty el aL 1983). These results
suggest that PCB sensitive phytoplankton experience a reduction in photosynthetic rates due to
irreversible damage to their chloroplasts. Geike and Parasher (1978) have shown that 5.0 ppm
24
-------
of HCB causes a 50 percent inhibition of photosynthesis in the alga Chlorella pyrenoidosa also
because of changes hi ultrastructure; 33.3 percent inhibition was noted at 0.1 ppm HCB and 42
percent at 1.0 ppm HCB.
Research on metals from atmospheric deposition and other sources has shown effects
including changes hi plankton community structure and significant decreases hi primary
production (Rybak el aL 1989). A 14 year study of a lake receiving waste from a heavy metal
mine uncovered the extinction or severe rarity of desmid and diatom species (Austin and
Munteanu 1984). Evidence therefore exists of possible perturbation of aquatic food chains
through substances other than pesticides or industrial chemicals.
Thus, the degree of effect of chemical exposure to marine and freshwater plankton is
highly dependent on species (due to natural variances hi sensitivity hi genotypes), chemical
mixture, and nutrient availability. Research indicates that pesticides and metals cause a reduction
hi primary production, however, this effect is usually temporary and does not occur at the
community level. A more important consequence of chemical exposure is the alteration of the
aquatic food chain, on a short- or long-term basis. The complete or partial loss of sensitive
species can cause a shift hi plankton community structure and composition which can potentially
alter an entire food chain, with repercussions which are yet undefined. Tinkering with the very
base of an ecosystem's food web could cause shifts hi predator/prey ratios and relationships
throughout trophic levels thereby changing the composition of food sources in the highest
echelons of the food chain. Although most of the studies described above were conducted with
concentrations higher than those presently recorded hi the environment, the absorption and uptake
of many of these chemicals by plants and live and dead plankton alike undermines the levels
recorded in water from streams and lakes.
Effects of these pollutants on humans and aquatic life are all considered to be from
chronic exposure. There are no known instances of acute toxic effects of these compounds in
any of the Great Water regions.
The populations at risk from exposure to these compounds include the top predators in
the aquatic food webs (e.g., sport fish); fish-eating wildlife (e.g., mink (Mustela vison), eagles,
gulls, terns, etc.); and human populations which consume large quantities of fish from Great
Waters areas (e.g., commercial fishermen and families, charter boat operators and families,
subsistence anglers such as Vietnamese, Native Americans) children, older people, and women
of childbearing age (concern for fetal exposure).
2.2.3.4 Biological Effects of Compounds of Concern
A number of chemicals transported atmospherically to water bodies are affecting the
health of wildlife and humans. Few of these chemicals are acute toxicants, powerful human
carcinogens, or genotoxicants at ambient concentrations (Colbom 1989). However, they are
developmental toxicants capable of altering the formation and function of critical physiological
25
-------
systems and organs. Thus, the developing embryo, fetus, and breast feeding offspring are
particularly sensitive to these chemicals (Table 8). This section summarizes the deleterious
effects of these contaminants on development, function, reproductive potential, behavior, and
disease processes in animals and humans as a result of exposure associated with freshwater and
marine resources. Each effect will be discussed in detail in the following sections covering the
discrete and multiple impacts of these compounds of concern.
Residues of the chemicals of concern have been reported in human tissue (Table 9),
including reproductive tissue (Table 10). For some of the chemicals an association has been
made between body burdens of the chemicals for those who regularly include fish in their diet
(Table 11). Mykkanen and coworkers (1986) estimated that 1 percent of daily energy, 1 percent
of daily cadmium, and 37 percent of daily mercury intake is from fish in the diet of Finnish
children.
Two of the atmospherically transported compounds of concern are not toxic substance,
but rather, are nutrients. Nitrogen and phosphorus are of concern because of their impacts on
the eutrophication of estuaries and freshwaters, respectively. The effects of these compounds will
be considered separately under the heading "Eutrophication". The effects of toxic compounds
will be discussed in the sections entitled "Cancer, "Immune System Impairment", "Metabolic
Impairment", "Nervous and Behavioral Impairment", "Endocrine Disruption", "Reproductive
Impairment" and "Transgenerational Effects".
Under ideal circumstances, an investigation into the quality of the data for each study
utilized in the preparation of a review manuscript would be made. Such data quality review is
obviously beyond the scope of this effort. However, a series of decisions made prior to the
inception of this project serve to establish relative confidence in the data used.
The studies and the information used in the preparation of the various sections of this
document are the most currently available data. Every effort has been made to restrict the use
of older studies to the role of comparison with contemporary data. In most cases, the older
studies have been utilized to either compare or contrast the older evidence with current
contributions and new knowledge. Further, efforts were made not to incorporate a single study
indicating a unique endpoint, and to present it hi the absence of supporting information.
Whenever possible, supporting studies have also been incorporated and discussed. In this
fashion, the question of individual data quality within a single study is minimized, and a relative
degree of confidence in the complete data set presented can be achieved.
26
-------
FIGURE 4
ROUTES OF TCDD EXPOSURE FOR HUMANS
Water &
Oroundwaier
.2 "9
Drinking Water L-J
~ Respiration'
1 "9
Almosnftoro
-------
TABLES
POPULATIONS AT RISK FROM EXPOSURE TO TOXIC POLLUTANTS
POPULATION AT RISK
piscivorous fish
fish-eating wildlife and birds
commercial fishermen
charter boat operators
subsistence fish eaters
children
elderly
women of childbearing age
28
-------
TABLE 9
EFFECTS OF COMPOUNDS OF CONCERN IN HUMANS
Compound
2,3,7,8-TCDD
(Dioxin)
Benzo[a]pyiene
(B[a]P)
Cadmium (Cd)
Chlordane
DDT/DDE
Dieldrin
i
Genotoxic
E
ATSDR
1987
0
ATSDR
1987
0
ATSDR
1987
0
Cabral
1985
0
ATSDR
1987
Carcinogenic
0
E
ATSDR 1987
Kazahtzis el aL
1988
0
IARC 1986
0
Falck el aL 1992
E
ATSDR 1987
Reproductive
Effects
0
0
0
0
0
0
Developmental
Effects
E
Erickson el aL 1984
0
E
Bonithon-Kopp el
aL1986a
0
0
0
Immunotoxic
E
ATSDR 1987
0
0
0
0
E
ATSDR 1987
Neurological
Effects
E
Barbieri el aL
1988,
Levy 1988
0
0
USEPA 1985,
ATSDR 1988
WHO 1979
ATSDR 1987
Target Organ
Damage
ATSDR 1987
0
ATSDR 1987
0
0
0
Accumulated
in Human
Tissues
Jensen 1987
0
Piscator 1985,
Subramanian
d aL 1985
Taguchi &
Yakushiji
1988
Williams el aL
1988,
Davies & Mes
1987
Williams el aL
1988
-------
TABLE 9 (Cont.)
Compound
HCB
Lead (Pb)
lindane
Mercury
(Hg)
Miiex
PCB
Toxaphdne
Genotoxlc
0
E
EPA, 1989
0
E
ATSDR 1988
0
0
E
WHO 1984
Carcinogenic
0
1ARC 1986
EPA, 1989
0
IARC 1986
0
USEPA-ODW
1987
0
E
ATSDR 1987
0
Reproductive
Effects
USEPA 1987
ATSDR 1988
EPA
1986, 1990
0
0
0
E
ATSDR 1987
0
Developmental
Effects
USEPA 1987
-f
ATSDR 1988
EPA
1986, 1990
0
E
Noidberg 1988
0
ATSDR 1987
0
Immunotoxlc
0
E
ATSDR 1988
EPA 1986, 1990
0
E
WHO 1976
0
E
Shigematsu el
aL 1978,
Chang el aL
1980
0
Neurological
Effects
USEPA 1987
ATSDR 1988
EPA
1986, 1990
0
WHO 1976
0
0
WHO 1984
Target Organ
Damage
USEPA 1987
ATSDR
EPA 1986, 1990
0
Noidbeig 1988,
Gnibb el aL 1987
0
0
0
Accumulated In
Human Tissues
Williams ct aL 1988
Subiamanian
etaL1985
Mes etaL1977,
Davies & Mes 1987
Subiamanian
daL1985
Williams ct aL 1988
Williams etaL
1988,
Humphrey 1983
0
Legend; 0 = No information A zero (0) does not necessarily mean there is no effect;
E = Equivocal it can also mean that no studies have been done.
+ = Positive results
- = Negative results
-------
TABLE 10
COMPOUNDS OF CONCERN FOUND IN HUMAN REPRODUCTIVE TISSUE
COMPOUND
Cadmium
Chlordane (HE)
DDE/DDT
Dieldrin
HCB
Hg
Lead
i
Lindane (g-HCH)
(a-HCH)
(b-HCH)
Miiex
PCB
2,3,7,8-TCDD
OVARIAN FOLLICLE
Baukloh el aL 1985,
Trapp el aL 1984
Trapp el aL 1984,
(DDT) Baukloh el aL 1985
*
Trapp et aL 1984,
Baukloh el aL 1985
Trapp et aL 1984
Trapp et aL 1984,
Baukloh et aL 1985
Trapp et aL 1984,
Baukloh et aL 1985
Trapp et aL 1984,
Baukloh et aL 1985
Trapp et aL 1984,
Baukloh el aL 1985
PLACENTA
Korpel et aL 1986
USPHS-ATSDR 1987
Ando el aL 1985
Courtney and Andrews 1985
Capelli and Minganti 1986
Kuhnert et aL 1981
Korpela et aL 1986
Kuhnert & Kuhnert 1988
Ando el aL 1985
TESTICLE
Szmcynski & Waliszewski 1981
Dougherty et aL 1980,
Szmcynski & Waliszewski 1981,
(DDE) Bush et aL 1986,
Schecter et aL 1989
Szmcynski & Waliszewski 1981,
Dougherty el aL 1980,
Bush et aL 1986,
Schecter et aL 1989
Szmcynski & Waliszewski 1981
Szmcynski & Waliszewski 1981
Szmcynski & Waliszewski 1981
Dougherty el aL 1980,
Bush et aL 1986,
Schecter el aL 1989
Schecter el aL 1992
-------
TABLE 11
RESIDUES REPORTED IN HUMANS THAT SHOW AN ASSOCIATION WITH
THOSE WHO REGULARLY INCLUDE FISH IN THEIR DIET
Chlordane
DDE
HCB.
Lead
Lindane
Mercury
2,3,7,8-TCDD
Mirex
OCS
PCB
Wariishi el aL 1986
Wisconsin DOH 1987,
Rogan el aL 1986a,
Kanja el aL 1986,
Bush fit aL 1984,
Noren 1983,
Kreiss el aL 1981
Noren 1983
Dabeka el aL 1986
Sloof and Matthijsen 1988
Langworth el aL 1988,
Wisconsin DOH 1987,
Mykkanen el aL 19861,
Lommel el aL 1985
Schecter el aL 1990
WHO 1984
Lommel el aL 1985
Jacobson and Jacobson 1988,
Wisconsin DOH 1987,
Humphrey 1985,
Bush el aL 1984,
Schwartz el aL 1983,
Jensen and Slorach 1991
1 Mykkanen el aL 1986. Estimated one percent of daily energy; one percent of
daily Cd; and 37 percent of daily Hg intake are from fish.
32
-------
22A Ecosystem Level Effects of Toxic Substances
The biological effects of pollution can occur at a variety of levels of biotic organization,
from the subcellular to whole populations and ecosystems. The science relating effects of toxic
substances across these biotic scales is not well developed, and it is often quite difficult to state
precisely how an effect on the physiology of an organism or on cellular processes will be
expressed (if at all) at the scale of populations or ecosystems. Often, scientists are unable to
predict with any certainty because population numbers may be controlled largely by processes
other than reproduction — such as the survival of fish larvae in the face of a high predation
pressure or the extent of energy flow to the fish population up the food web. This does not
imply that populations and ecosystems are better buffered against the effects of toxic substances
than are lower levels of biotic organization (cells, organs, organisms), rather it suggests only that
there is great uncertainty in understanding the relationships among levels.
The effects of toxic substances on populations and ecosystems have received far less study
than have effects on individual organisms. However, recent reviews (Schindler 1987; Howarth
1991) have reached some general conclusions: changes in the structure of a community are a
more sensitive indicator of toxic stress than are changes in ecological processes such as primary
production; indirect effects resulting from subtle changes in competition and food web structure
can have major ramifications on populations and aquatic ecosystems; and unexpected effects from
pollution are commonly found in pollution studies.
Two examples can illustrate the complexity of the response of aquatic ecosystems to
stress. Whole-lake experiments at Canada's experimental lakes area showed that the major effect
of acidification on fish is an indirect one. While extreme acidification in these experiments
resulted in loss of trout without mobilization of aluminum by altering the structure of the food
web. The trout gradually starved and were unable to reproduce (Schindler si aL 1985).
In another example, an oil spill in the Baltic Sea resulted in decreased hatching success
of herring eggs, but the effect was not a result of direct toxicity on the eggs. Laboratory studies
showed a high tolerance of these fish eggs to oil. Rather the effect of the oil was to kill off
benthic amphipods, and the loss of the amphipods resulted in a fungal overgrowth of the fish
eggs, killing many of them. Normally, the amphipods graze upon the fungi and keep it under
control (Nellbring si aL 1980).
Thus, the state of present knowledge of the effects of toxic substances at the ecosystem
level is inadequate. Future research efforts will be required to enable an understanding of the
potential alterations in relationships among the various levels of ecosystem organization.
33
-------
2.2.5 Discrete Effects of Contaminants of Concern
2.2.5.1 Eutrophication
Eutrophication was recognized as a major problem in the Great Lakes and many estuaries
at least 30 years ago (Ryther 1954; Davis 1964; Beeton 1965; Ryther and Dunstan 1971; E.P.A.
1971). During the 1970's, management steps were taken to reduce the inputs of phosphorus to
the Great Lakes. As a result, Lakes Erie and Ontario have substantially recovered from
eutrophication (DePinto 1986; Lean 1987; Schindler 1987; DePinto 1991; Schelske and Hodell
1991). There has also been progress hi reducing eutrophication hi some limited estuarine areas
as well, such as coastal ponds on Long Island which were affected by runoff from duck farms
in the 1950's (Ryther 1989) and Kaneohe Bay hi Hawaii which received large sewage inputs until
the mid 1970's (Smith 1981). However, hi general, the problem of eutrophication hi estuaries
has grown (Office el aL 1984; Larsson el aL 1985; Rosenberg 1985; D*Elia 1987; Baden el aL
1990; Parker and O'Reilly 1991; Lein and Ivanov 1992; Jaworski el aL 1992). Recently,
eutrophication was identified as the most serious pollution problem facing the estuarine waters
of the United States (NRC 1993).
The principal reason for the slower remediation of estuarine waters is that, while
eutrophication hi lakes is controlled by phosphorus, nitrogen controls eutrophication in most
temperate-zone estuaries. More effort has been expanded to control phosphorus, and the sources
of nitrogen are more diffuse and difficult to control (Butt 1992). As a result, many estuaries
receive nitrogen inputs per unit area which are more than 1,000-fold greater than those of heavily
fertilized agricultural fields (Nixon el aL 1986). In moderation, nitrogen inputs to estuaries and
coastal seas can be considered beneficial, since they result in increased production of
phytoplankton (the microscopic algae floating hi water), which, hi turn, can lead to increased
production of fish and shellfish (Nixon 1988; 'Rosenberg el aL 1990; Hansson and Rudstam
1990). Excess nitrogen can be highly damaging, leading to effects such as anoxia and hypoxia
from eutrophication, nuisance algal blooms, dieback of seagrasses and corals, and reduced
populations offish and shellfish (Ryther 1954,1989; Kirkman 1976; McComb elaL 1981; Kemp
el aL 1983; Officer el aL 1984; Gray and Paasche 1984; Cambridge and McComb 1984; Larsson
el aL 1985; Price el aL 1985; Rosenberg 1985; D'Elia 1987; Rosenberg el aL 1990; Cederwall
and Elmgren 1990; Baden el aL 1990; Hansson and Rudstam 1990; Parker and O'Reilly 1991;
Lein and Ivanov 1992; Smayda 1992). Eutrophication also may change the plankton-based food
web from one based on diatoms toward one based on flagellates or other phytoplankton which
are less desirable as food to organisms at higher trophic levels (Doering el aL 1989).
In most estuaries, the sources of nitrogen are only poorly known. However, atmospheric
sources can be important, in sharp contrast to phosphorus inputs, for which air borne pathways
are generally quite minor (Wolfe el aL 1991; Jaworski el aL 1992). Inputs of nitrate and
ammonium directly to the surface waters of Long Island Sound from the atmosphere are
estimated to be at least 10 percent of the total nitrogen inputs (Wolfe el aL 1991). However,
indirect inputs of nitrogen from airborne sources are probably much larger, since over half of the
nitrogen comes from upstream sources and urban runoff (Wolfe el aL 1991). Studies of the
34
-------
watersheds of the entire Chesapeake Bay (Fisher and Oppenheimer 1991) and of the upper
Potomac River (Jaworski el aL 1992) have suggested that 28 percent and 40 percent, respectively,
of the nitrogen fluxes into the watershed come from atmospheric deposition. Not all of the
nitrogen deposited on a watershed flows downstream to an estuary; studies in several watersheds
near Chesapeake Bay have suggested that roughly two thirds of the nitrogen deposition falling
on forested lands is retained in the forest (Groffman and Jaworski 1991; Jaworski el aL 1991).
The factors controlling nitrogen retention by forests are poorly known, but uptake by trees is
probably a major mechanism (Jaworski el aL 1991) since many forests are nitrogen limited
(Vitousek and Howarth 1991). However, fully mature forests presumably will not retain as much
nitrogen because there is no net growth of trees (Jaworski el aL 1991). Further, if sufficient
nitrogen is added to a forest via deposition, the forest can become nitrogen "saturated" (Aber el
aL 1991). Increasing concentrations of nitrate in streams in the Catskill Mountains of New York
over the past decade suggest that the forests there have become saturated and are now exporting
more nitrogen downstream (Murcoh and Stoddard 1991).
Nutrient Limitation
Nitrogen and phosphorus are essential nutrients for plant growth. Phytoplankton
production in most lakes, coastal marine ecosystems, and estuaries is nutrient limited. As a
result, increased nutrient inputs lead to higher production and eutrophication (Edmondson 1970;
Ryther and Dunstan 1971; Vollenwieder 1976; Schindler 1977, 1978; Schindler el aL 1978;
Graneli 1978, 1981, 1984; McComb el aL 1981; Boynton el aL 1982; Nixon and Pilson 1983;
Wetzel 1983; Valiela 1984; Smith 1984; Nixon el aL 1986; DElia el aL 1986; D"Elia 1987;
Howarth 1988; Andersen el aL 1991). Unfortunately, the discussion of nutrient limitation in
aquatic ecosystems has been surrounded by some confusion, in part because the term can have
many different meanings and is often used quite loosely (Howarth 1988). Further, potential
methodological problems in determining nutrient limitation increase the confusion (Hecky and
Kilham 1988; Howarth 1988; Banse 1990). In the case of eutrophication, the appropriate
definition of nutrient limitation is the regulation of the potential rate of net primary production
by phytoplankton (Howarth 1988). Net primary production is defined as the total amount of
photosynthesis minus the amount of plant respiration occurring in a given area (or volume) of
water in a given amount of time. If an addition of nutrients would increase the rate of net
primary production — even if such an addition means a complete change in the species
composition of the phytoplankton, production is considered to be nutrient limited (Howarth 1988;
Vitousek and Howarth 1991).
Factors other than nutrient input can also influence or partially control primary production.
For example, phytoplankton production in some estuaries (e.g., the Hudson River) is limited by
light availability. Such light limitation tends to occur in extremely turbid estuaries, or in
estuaries which moderate turbidity coexists with deep mixing of the water. The turbidity can
result both from suspension of inorganic particles and from high phytoplankton biomass. Thus,
light limitation often is a result of self-shading by the phytoplankton (Wetzel 1983). In estuaries
where nutrient inputs are high and production is limited by light, the nutrients are simply
transported further away from the source before being assimilated by phytoplankton, e.g., the
35
-------
Hudson River and New York Harbor into the New York Bight (Malone 1982). This transport
may or may not provide sufficient dilution to avoid excessive eutrophication. Frequently,
eutrophication simply occurs further afield from the nutrient source.
Zooplankton and other animals can influence the rate of primary production and the
biomass of phytoplankton by their grazing on phytoplankton. This phenomenon has received
extensive study and discussion in both freshwater ecosystems (Carpenter el aL 1985,1987; Morin
el aL 1991), and hi offshore ocean ecosystems (Steele 1974; Banse 1990). However, the effects
of grazing are largely unstudied in estuaries and coastal seas (Rudstam el aL 1992). In lakes,
higher abundances of phytoplankton and higher rates of net primary production occur when
zooplankton biomass is lower (Carpenter el aL 1987; Morin el aL 1991). Changes in the
abundance and species composition of fish (Carpenter el aL 1985) and of filter-feeding benthic
organisms may also affect phytoplankton abundance. For instance, water clarity in Lake Erie has
increased greatly after the unintentional introduction of zebra mussels (E. Mills 1992, personal
communication). In general nutrient supply should be viewed as the cause of eutrophication,
with grazing pressures being a secondary regulator.
Nitrogen Versus Phosphorus Limitation
In the 1960's and early 1970's, there was intense debate over which nutrient controlled
eutrophication in lakes (see papers in the volume edited by Likens 1972). By the late 1970's,
however, phosphorus inputs were clearly identified as the major factor, at least in mesotrophic
and eutrophic lakes (Vollenwieder 1976; Schindler 1977, 1978; Schindler el aL 1978; Wetzel
1983). As a result, management strategies were undertaken to reduce phosphorus inputs into the
Great Lakes. These strategies have been successful and, in response, these lakes recovered from
eutrophication during the 1980's (DePinto 1986; Lean 1987; Schindler 1987; DePinto 1991;
Schelske and Hodell 1991).
In contrast to the Great Lakes and most other temperate-zone lakes, nitrogen is probably
the element usually limiting to primary production by phytoplankton in most estuaries and coastal
seas of the temperate zone (Ryther and Dunstan 1971; Vince and Valiela 1973; Smayda 1974;
Norm 1977; Graneli 1978, 1981, 1984; Boynton el aL 1982; Nixon and Pilson 1983; Valiela
1984; Nixon el aL 1986; D'Elia el aL 1986; Howarth 1988; Frithsen el aL 1988; Rydberg el aL
1990; Vitousek and Howarth 1991; Nixon 1992). However, some temperate estuaries such as the
Apalichicola in the Gulf of Mexico may be phosphorus limited (Myers and Iverson 1981;
Howarth 1988) and others, e.g., parts of Chesapeake Bay and the Baltic Sea, may switch
seasonally between nitrogen and phosphorus limitation (McComb el aL 1981; D'Elia el aL 1986;
Graneli el aL 1990; Andersen el aL 1991). Many tropical estuarine lagoons also may be
phosphorus limited (Smith 1984; Smith and Atkinson 1984; Howarth 1988; Vitousek and
Howarth 1991).
The question of nitrogen limitation of primary production hi most temperate-zone
estuaries and coastal seas was much debated throughout the 1980's (D'Elia 1987; Howarth 1988;
Nixon 1992). One argument against nitrogen limitation was that phosphorus is generally limiting
36
-------
in temperate-zone lakes and, until recently, there was little evidence that the biogeochemical
processes regulating nutrient limitation were fundamentally different in freshwater as compared
with marine ecosystems (Schindler 1981; Smith 1984). Another argument was that much of the
evidence for nitrogen limitation in marine ecosystems came from extremely short-term, small-
scale enrichment experiments in flasks or bottles. It may not be possible to extrapolate the
results of such short-term enrichment experiments to an entire ecosystem (Smith 1984; Hecky
and Kilham 1988; Howarth 1988; Marino el aL 1990; Banse 1990).
In recent years, increasing evidence has accumulated indicating that nitrogen is limiting
in many coastal marine ecosystems, and that the biogeochemical processes regulating nutrient
limitation do vary between marine and freshwater ecosystems. The new evidence for nitrogen
limitation consists of generally low concentrations of dissolved nitrogen compared with dissolved
phosphorus (Boynton elaL 1982; Graneli 1984; Valiela 1984) and longer, large-scale enrichment
experiments (D'Elia el aL 1986), including one mesocosm experiment of many months duration
(Frithsen el aL 1988; Nixon 1992; Frithsen el aL, unpublished data). While any one such piece
of evidence may not be entirely convincing, the good agreement among the several studies
convincingly demonstrates nitrogen limitation (Howarth 1988; Vitousek and Howarth 1991).
At least three factors in the biogeochemical cycles appear important to the question of
nitrogen or phosphorus limitation: (1) the ratio of nitrogen to phosphorus in nutrient inputs to
estuaries is frequently less than for lakes, (2) the sediments are often a more important sink of
phosphorus in lakes than in marine ecosystems, and (3) nitrogen fixation is a more prevalent
process in the plankton of lakes (Howarth 1988). Each of these differences is discussed briefly
below.
(1) In both freshwater and marine ecosystems, the relative requirements of
phytoplankton for nitrogen and phosphorus are fairly constant, with the two elements
being assimilated in the approximate molar ratio of 16:1 (Redfield 1958). If there were
no biogeochemical processes acting within a water body, the ratio of nitrogen to
phosphorus in the nutrient inputs to the ecosystem would determine whether the system
were nitrogen or phosphorus limited, with ratios below 16:1 leading to nitrogen limitation
and higher ratios leading to phosphorus limitation (Howarth 1988). In fact, the N:P ratios
in nutrient loadings to many estuaries and coastal seas are below this ratio, while nutrient
inputs to temperate lakes tend to have higher N:P ratios (Jaworski 1981; Kelly and Levin
1986; NOAA/EPA 1988). This difference hi ratios probably reflects the relative
importance of sewage, which tends to have a low N:P ratio, as a nutrient source of coastal
waters.
(2) Biogeochemical processes within sediments act to alter the relative abundance of
nitrogen and phosphorus in an ecosystem. Denitrification, the bacterial reduction of
nitrate to molecular nitrogen, removes nitrogen and tends to make coastal marine
ecosystems more nitrogen limited (Nixon el aL 1980; Nixon and Pilson 1983). However,
this process appears to be even more important in lakes than hi estuaries and coastal seas;
a higher percentage of the nitrogen mineralized during decomposition is denitrified in lake
37
-------
sediments than in estuarine sediments (Seitzinger 1988; Gardner el aL 1991; Seitzinger
el aL 1991). Of more importance in explaining a tendency for nitrogen limitation in
coastal marine ecosystems of the temperate zone, therefore, is the relatively high
phosphorus flux from sediments; nutrient fluxes from these sediments have fairly low N:P
ratios (Rowe el aL 1975; Boynton el aL 1980; Nixon el aL 1980). In many lakes,
phosphorus is bound in the sediments (Schindler el aL 1977), although in others,
phosphorus fluxes are comparable to marine sediments (Khalid el aL 1977). Nutrient
fluxes from lake sediments can be either enriched or depleted in nitrogen relative to
phosphorus (Kamp-Nielsen 1974). Caraco el aL (1989, 1990) have suggested that the
abundance of sulfate hi an ecosystem partially regulates the sediment flux of phosphorus,
with phosphorus binding in sediments being greatest where sulfate concentrations are
lowest. This suggestion is consistent with variable fluxes hi lakes and higher fluxes in
coastal marine ecosystems.
(3) When the relative abundance of nitrogen to phosphorus is low in the water column
of lakes, nitrogen-fixing species of cyanobacteria are favored since they can convert
molecular nitrogen to ammonium or organic nitrogen. Under such nitrogen-depleted
conditions in lakes, these cyanobacteria often are the dominant phytoplankton species and
fix appreciable quantities of nitrogen. As a result, nitrogen deficits (relative to
phosphorus) can be alleviated, and primary production hi the lake is phosphorus limited
(Schindler 1977; Flett el aL 1980; Howarth 1988; Howarth el aL 1988a). In contrast,
nitrogen-fixing cyanobacteria are rare or absent from the plankton of most estuaries and
coastal seas, a condition helping to maintain nitrogen limitation hi these ecosystems
(Howarth 1988; Howarth el aL 1988a). Exceptions are found hi the Baltic Sea (Lindahl
and Wallstrom 1985) and in the Australian Harvey-Peel estuary (McComb el aL 1981),
but are unknown in the waters of the U.S. The explanation for the rarity of planktonic,
nitrogen-fixing cyanobacteria in coastal marine waters is still subject to debate (Howarth
el aL 1988b; Paerl el aL 1987; Paerl and Carlton 1988; Carpenter el aL 1990; Marino el
aL 1993). Possible reasons include one or more of the following: a lower availability
of iron and molybdenum required for nitrogen fixation in saline water (Howarth and Cole
1985; Howarth el aL 1988b; Marino el aL 1990), greater turbulence in coastal marine
systems, allowing oxygen to poison the nitrogenase enzyme responsible for nitrogen
fixation (Paerl el aL 1987; Paerl and Carlton 1988); greater grazing pressure on
cyanobacteria in marine systems (Vitousek and Howarth 1991); and a lower light
availability hi estuaries and coastal waters due to higher turbidity and/or deeper mixed
layers (Howarth and Marino 1990; Vitousek and Howarth 1991).
As noted above, many tropical estuaries and coastal systems may be phosphorus limited
(Smith 1984; Smith and Atkinson 1984). Although the evidence for limitation of production by
phytoplankton is not entirely clear in tropical systems (Howarth 1988), and production by
seagrasses and attached macroalgae is sometimes nitrogen limited in tropical systems (Lapointe
el aL 1987; McGlathery el aL 1992), primary production by seagrasses in many tropical areas is
clearly limited by phosphorus (Short el aL 1985; 1990; Littler el aL 1988; Powell el aL 1989).
Phosphorus limitation in these systems is probably the result both of a high degree of phosphorus
38
-------
adsorption in the calcium-carbonate sediments which dominate such tropical systems (Morse el
aL 1985) and the high rates of nitrogen fixation associated with benthic algal mats and with
symbionts of seagrasses in clear, relatively oligotrophic lagoons (Howarth 1988; Howarth el aL
1988a).
2.2.5.2 Cancer
None of the airborne compounds of concern are documented carcinogens in humans at
ambient concentrations. However, occupational exposure to cadmium (Kazantzis el aL 1988),
dioxin (Fingerhut el aL 1991; Manz el aL 1991) and B(a)P (ATSDR 1987) has been correlated
with cancer. Falck el aL (1992) found elevated levels of PCB, DDT, and DDE hi fatty breast
tissue from women with breast cancer compared with breast tissue from women with non-
malignant breast disease.
Other than reports on dermal and liver cancers in fishes and the beluga whales
(Delphinapterus leucas) in the St. Lawrence River and Estuary, reports of cancer in wildlife are
rare. In each of these cases the causal agents were discovered to be polyaromatic hydrocarbons
(PAHs) hi follow-up laboratory studies (Black el aL 1981; Black el aL 1982; Baumann and
Harshbarger 1985; Hayes el aL 1987; Cairns and Fitzsimmons 1987; NOAA 1991).
High incidences of liver neoplasms in fish from highly contaminated sites hi Puget Sound,
Washington, have been reported along with assorted preneoplastic and regenerative lesions in
English sole (Parophrys vetulus), rock sole (Lepidopsetta bilineata), and starry flounder
(Platichthys stellatus) (NOAA 1991). Field and laboratory studies linked contaminant exposure
not only to the liver neoplasms/lesions, but also to other metabolic effects. Sediments and PAHs
extracted from sediments from contaminated harbors applied dermally and fed to fish induced
dose-related tumors in the confined fish. Other fish species exhibiting similar lesions include
the black croaker, flathead sole (Hippoglossoides elassodon), hardhead catfish, white croaker
(Genyonemus lineatus), white perch (Morone americana), windowpane flounder, and winter
flounder (Pseudopleuronectes americanus) (NOAA 1991).
Follow-up long-term field studies at other US locations supported the Puget Sound
findings (Varanasi 1989). A high prevalence of liver lesions and/or neoplasms was found in
starry flounder, black croaker, and winter flounder hi San Francisco Bay, the Oakland Estuary,
San Diego Bay, and the North East coast, respectively. Boston Harbor, East Raritan Bay, and
Salem Harbor, all contaminated with aromatic hydrocarbons and PCBs, had whiter flounder with
high liver contaminant concentrations associated with liver neoplasms. Great Lakes studies
revealed that epidermal papillomas, liver lesions, and a tumor were induced by topical or dietary
exposure of bullheads to Buffalo River and Black River sediments (MacCubbin el aL 1985;
Baumann el aL 1987; Black el aL 1985).
In the Chesapeake Bay ecosystem, liver neoplasms and other lesions were found in the
mummichog (FundUlus heteroclitus) from Elizabeth River sites (Vogelbein el aL 1990) and 15
39
-------
percent of the white perch from 15 estuarine tributaries (May el aL 1988). Ninety-three percent
of the fish from the contaminated Elizabeth River site had visible hepatic lesions; thirty-three
percent had hepatic carcinomas. Vacuolized liver cells were found in striped bass (Morone
saxatilis) and other fish of the Choptank River, the Chesapeake and Delaware (C&D) Canal, the
Potomac River near Quantico, and upper bay at the Susquehenna (Hall el aL 1987, 1988a, b).
In addition, renal lesions were found in increased frequencies in Elizabeth River fish
CThiyagarajah el aL 1989) and in yearling striped bass from the Potomac River (Hall el aL 1987).
Gill hypertrophy and gill lesions were also found hi fish species exposed to water from the
Elizabeth River, C&D canal sites, and the Potomac River (Hargis and Zwerner 1988; Hall el aL
1987; Hall el aL 1988b). Further, cataracts hi spot, Atlantic croakers (Micropogonias wuhdatus),
weakfish, spotted hake, and gizzard shad, as well as fin erosion in toadfish were attributed to
benzo-a-pyrene in the Elizabeth River (Hargis and Zwerner 1988; Huggett el aL 1987).
At the organismic level, populations of commercially and ecologically valuable fish
species which spawn in the Chesapeake Bay watershed are declining, suggesting an
environmental impact which affects the spawning grounds (fresh-water and tributaries) (Wright
61 aL 1992). The health disturbances exhibited by fish species of the Chesapeake Bay estuary
cannot be correlated directly to any one chemical or heavy metal in the natural environment
(Helz and Huggett 1987; Wright el aL 1992). Wright and coworkers (1992) analyzed patterns
of similarity for acute and sublethal effects across species and found that, of the heavy metals,
copper and mercury were the most acutely and chronically toxic; and that insecticides were of
greater detriment to aquatic organisms than herbicides. PAHs in the Elizabeth River, as with the
Puget Sound studies on the English sole, contribute to the observed neoplasms in fish (Wright
el aL 1992). Direct correlation between toxic chemicals and metals and the health effects
observed in the Chesapeake Bay wildlife remains incomplete due to limited information at the
population and community level; interaction of physical conditions such as salinity, pH, and
temperature; the presence of disease organisms; and predation, competition, and human
involvement hi population survival (Wright el aL 1992). However, the prevalence of health
disturbances, the loss of species diversity hi the Bay, and the gradient of effects matched with
the gradient of contamination from urban to remote sites indicate a contribution to the effects
from toxic chemicals (Wright el aL 1992).
A 40 percent incidence of tumors was discovered in stranded beluga whales
(Delphinapterus leucas) hi the St. Lawrence necropsied between 1983 and 1990 (Beland el aL
1992; Martineau el aL 1988, 1987, 1985). Although these studies were performed on dead
animals, age distribution studies confirmed that they were representative of the live population.
The tumors found in the 1987-1990 group affected multiple organs (mammary, pulmonary,
intestinal, gastric, and thymus) and were reported as malignant, benign, and abdominal mass.
Over a ten year period 46 percent of the belugas had at least one tumor (Beland el aL 1992).
The chemical contaminant.levels of the St. Lawrence belugas were significantly higher than in
Arctic belugas for mercury, lead, total DDT, PCBs, and mirex. Benzo-a-pyrene (BaP) DNA-
adducts hi brains and livers were discovered in 8 of 9 belugas tested (Beland el aL 1992).
40
-------
The following stages of carcinogenesis in fish have been described: (1) initiation of
tumorigenesis through exposure to known carcinogens such as B(a)P found in sediments and
suspended in the water column; (2) promotion of tumorigenesis by PCBs on initiated cells; and
(3) decreased immune function resulting from concomitant exposure to organochlorine
contaminants that are known immune suppressants (Black el aL 1981; 1982; Baumann and
Harshberger 1985; Hayes el aL 1987; Cairns and Fitzsimmons 1987).
2.2.5.3 Immune System Impairment
Linking immune system impairment with exposure to a toxic chemical(s) has been
confounded by the presence of natural agents such as viruses and other pathogens which exhibit
comparable symptoms in humans and wildlife. Although a direct cause-effect linkage has not
been established with regard to immune suppression and xenobiotics hi wildlife, a body of
evidence exists in laboratory studies which demonstrate xenobiotic effects on the immune system.
This section presents field observations of reduced immunocompetence in animals carrying
elevated contaminant body burdens. Laboratory evidence of immunological changes in the
presence of the same contaminants is also presented.
Wildlife Studies
Since 1987, an increased number of marine mammal mortality events and strandings have
occurred in the Northern Hemisphere (Table 12). Dead or dying seals, dolphins, porpoises, and
whales have been observed from the Pacific Northwest to the eastern coast of the U.S., the Gulf
of Mexico, the Mediterranean Sea, and the Baltic and North Seas (Geraci 1989; Harwood el aL
1989; Lavigne and Schmitz 1990; Kuehl sL aL 1991; Raga and Aguilar 1991; UNEP 1991;
Sarokin and Schulkin 1992). General systemic infections, organ lesions, poor health, and
inability to combat infection characterized animals in the die-offs. Factors suspected of
contributing to the cause of death included newly discovered viral agents, called morbilliviruses,
similar to canine distemper that are specific to seals or dolphins (Kuehl el aL 1991); climatic
change resulting in a warmer environment conducive to the spread of contagious agents (Lavigne
and Schmitz 1990); algal blooms producing neurotoxins, such as brevitoxin from red tide (Geraci
1989); and increased body levels of organohalogens (Raga and Aguilar 1991). Bottlenose
dolphins from the Atlantic coast and striped dolphins from the Mediterranean Sea had liver, lung,
and lymphatic system lesions. The liver lesions hi striped dolphins and depleted lymphocyte
follicles hi bottlenose dolphins suggested chemical immunosuppression (Borrell and Aguilar
1991). In either case, the lesions could not be attributed to viral infection. Immunotoxic
environmental agents were also cited as a possible cause of lymphoid depletion hi pinnipeds on
the southern California coast (Simpson and Gardner 1972; Cavagnelo 1979; Britt and Howard
1983). It is important to note that all of the affected marine species are toothed and dependent
upon fish.
A ten year monitoring program revealed that the troubled population of beluga whales at
the mouth of .the St. Lawrence River hold significantly higher body burdens of PCBs, DDT, and
41
-------
mirex than other declining marine mammal populations and the least contaminated, healthy
population of Arctic beluga whales (Beland el aL 1991). Researchers suggested that general poor
health, susceptibility to bacterial and viral infections, tumors, and other pathological abnormalities
within the St. Lawrence population were the result of immunosuppressive activity of
environmental contamination origin (Martineau el aL 1987; Muir el aL 1990; Beland el aL 1992).
Beland (1992) determined that American eels are the vector for 100 percent of the mirex, 37
percent of the PCBs, 15 percent of the DDT hi the St. Lawrence belugas. The migrating eels
transport the material as they return from the Great Lakes to the Atlantic Ocean to spawn.
European field researchers tested the association between organochlorine chemicals and
population decline in the harbor seal (Phoca vitulina) (Reijnders 1986; Brouwer el aL 1989).
They found an association between PCBs and DDT and reproductive loss (see Section 2.2.5.7)
and immune system function.
In the Chesapeake Bay ecosystem, biota have experienced similar impacts on their
immune systems. Diminished immune response was demonstrated by decreased macrophage
phagocytic activity hi bottom-dwelling fish species of the Elizabeth River as compared with the
York River (Warriner el aL 1988; Weeks and Warriner 1984; Weeks el aL 1986).
Saxena el aL (1992) found significant decreases in catfish (Heteropneustis fossilis)
humoral immune response to the microorganism Aeromonas hydrophila resulted from low-level
exposure to cadmium and hexachlorocyclohexane (HCH). Antibody titre, erythrocyte count,
leukocyte count, hemoglobin, hematocrit, and total plasma protein were reduced significantly by
the combination of HCH and cadmium. HCH and cadmium alone resulted in a significant
reduction of erythrocytes, leukocytes, and hemoglobin. The effect seen with a combined
exposure to cadmium and HCH indicated a synergistic immunosuppressive chemical action.
Erdman (1983) found evidence of immune system impairment hi Forster's and common terns
(Sterna hirundo) experiencing a post-fledgling die-off hi 1988.
Laboratory And Mechanistic Studies
The immune system is characterized by a highly responsive and integrated system of cells
and tissues. The integrated nature of the immune system complicates and magnifies the effects
of xenobiotics. The impairment of certain cells (such as helper T-cells) subsequently disrupts
the function of other cells, such as cytotoxic T-cells and antibody-producing B-cells. The
mechanism of immune-response impairment is best understood in the case of TCDD, although
many of the effects of PCB are similar, and may operate through a similar mechanism.
Relationships between sublethal exposure to PCBs, DDT, dieldrin, and dioxin and immune
system dysfunction are substantiated by experimental studies (Tables 13 and 14).
Observations of significant impairment in both the cellular and humoral immune response
to the chemicals of concern are as follows:
42
-------
susceptibility to viral and/or bacterial infection
reduced antibody synthesis
complement synthesis compromise
thymic atrophy
lymphoid depletion
decreased macrophage, phagocyte, and bactericidal activity
suppressed IgM response in offspring from maternal exposure.
TCDD is a potent immunosuppressant in laboratory animals (Sonawane el aL 1988;
Holsapple el aL 1991). Effects include changes in innate and acquired immunity, including both
humoral (antibody) and cell-mediated immune responses (Holsapple el aL 1991; Morris el aL
1991). The ED50 for suppression of plaque-forming cells (immunosuppression) of TCDD is 2.4
nmol/kg, and that of 2,3,4,7,8-PCDF, the most persistent and predominant congener found in
human tissues, is 3.0 nmol/kg (Davis and Safe 1988).
Central to the immunosuppressive effects of xenobiotics are their effects on the major
immune cell producing organs, the thymus and spleen. Reduction in thymic weight begins 4
days following administration of TCDD (Gorski el aL 1988), and will lead to eventual depletion
of mature lymphocytes (Ivans el aL 1992). In birds, TCDD-induced immunodeficiency occurs
by reducing the number of lymphoid cells in the bursa of Fabricius in a dose dependent manner
(Nikolaidis el aL 1988).
43
-------
TABLE 12
MAJOR MARINE MAMMAL DIEOFFS
COMMON NAME SPECIES
Dolphin, bottlenose Tursiops truncates
Dolphin, striped Stenella coeruleoalba
Seal, Baikal
Seal, grey
Seal, harbor
Seal, ringed
Whale, beluga
Whale, humpback
Whale, sperm
Phoca sibirica
Halichoerus grypus
Phoca vitulina
Phoca hispada
Delphinapterus leucas
Megaptera novaeangiea
Physter macrocephalus
YEAR
TDCATTON
1987-1991 Eastern Coast, Australia
1987-1988 North Atlantic, U.S.
1990
1992
Gulf of Mexico, U.S.
Matagora Bay, TX, U.S.
1990-1992 Mediterranean Sea
1987-1988 Lake Baikal, Siberia
1987-1988 Baltic & North Seas, Europe
1987-1988 Baltic & North Seas, Europe
1987-1988 Baltic & North Seas, Europe
1979-1992 St. Lawrence Estuary, Canada
1987 North Atlantic, U.S.
1988-1990 European/Norwegian Coasts
CITATION
Dayton 1991,
Geraci 1989,
Kuehl el aL 1991
Lancaster 1990,
Potter 1992
Raga and Aguilar
1992
Simmonds 1991,
UNEP 1991
Harwood el aL 1989
Dietz el aL 1989,
Addison 1989
Oehme el aL 1990
Beland el aL 1992
Geraci 1989
Simmonds 1991
-------
TABLE 13 (Cent.)
COMPOUND
B[a]P
Cadmium
Chlordane
Chlordane
DDT/DDE
CITATION
Bozelka and Salvaggio 1985
USPHS-ATSDR 1988
Myers el aL 1988
Bozelka and Salvaggio 1985
USPHS-ATSDR 1988
Blakley 1988
Cifone el aL 1989
Bozelka and Salvaggio 1985
Barnett el aL 1985*
Beggs el aL 1985
Johnson el aL 1987
Blaylock el aL 1990
Menconi el aL 1988
Kaminski el aL 1986
Banerjee el aL 1986, 1987 a, b
EFFECT
5 ng/ml to human lymphocytes dose dependent inhibition cell
proliferation. Inhibits IL2 production and partially receptor
expression.
-
CTL and NK responses differ in adult offspring of mice
peanut butter prenatally 0, 4, or 8 mg/day/b.w depending
age and sex.
fed
on
<1 (ig/m3 to > 5 |xg/m3 dose-response relationship with
sinusitis, bronchitis, and migraine in residents in homes
treated.
Macrophages in vitro exhibited significant phagocytotic
ability.
Altered cell-mediated responses, decreased IgM-antibody
production in rodents.
-------
TABLE 13
TOXIC SUBSTANCES AFFECTING AN ALTERATION IN IMMUNE FUNCTION IN VIVO AND IN VITRO
COMPOUND
2,3,7,8-TCDD
2,3,7,8-TCDD
Aldicarb
CITATION
Sonawane el aL 1988
Jennings el aL 1988
d'Argy at al. 1989
McConkey and Orrenius 1989
Gorski el aL 1988
Davis and Safe 1988
Davis and Safe 1989
Fine el aL 1988
Luster el aL 1988.
Spitzbergen el aL 1988
Selvan el aL 1989
EFFECT
Increased corticosterone at 25 fig/kg in S-D rats, decreased
thymus weight, morphological changes in thymus & adrenal
over starvation stress.
25 mg/kg A1254 with 3.7 nmol/kg TCDD (immunotoxic
dose) reduced TCDD toxicity.
A1260, 1254, 1248, 1242, 1016, & 1232 ED50 to inhibit
SRBC is 104, 118, 190, 391, 408, & 464 mg/kg or 0.28,
0.35, 0.66, 1.5, 1.5 & 2.0 nmol/kg, respectively.
Reconstituted breast milk congeners required 50 mg/kg to
antagonize 3.7 nmol TCDD.
Maternal single dose 10 jig/kg led to TdT 70-90 percent
inhibition in 4-11 day-old mice bone marrow. Thymic
[TCDD] 1-31 fg/mg tissue.
2 |ig/kg elicits T-dependent and T-independent antibody
response in vivo and ED50 7 nM after in vitro additions to
spleen culture.
1 fig/kg caused decrease in lymphoid cells in thymus, splenic
lymphoid depletion, hypocellularity of blood forming tissues
in rainbow trout.
Suppresses macrophage-mediated cytotoxicity of tumor cells
at 0.1 ppb i.p (f) C3H mice.
-------
TABLE 13 (Cont.)
COMPOUND
Mirex
PCB
(See Table 15)
PCP
TBT
TBT
Toxaphene
CITATION
WHO 1984
USPHS-ATSDR 1988
Shigematsu el aL 1978
Smialowicz el aL 1989
Smialowicz el aL 1989
Bozelka and Salvaggio 1985
WHO 1980
Snoeji 1987
Snoeji 1988
Smialowicz el aL 1989
Smialowicz el aL 1989
Van Loveren el aL 1990
Bozelka and Salvaggio 1985
EFFECT
10 & 25 mg/kg after 15 week male Fischer 344 rats, thymic
involution & NK cell activity and LP response only at 25
mg/kg.
Hepatomegaly at 1 mg/kg and thymic invol at 10 mg/kg after
5 weeks.
Dose causing 50 percent reduction in thymus weight was 18
mg DBTC & 29 mg TBTC/kg bw rats.
2.5 mg/kg x 10 produced thymic invol. & mitogen response
suppressed at 5 mg/kg, adult male Fischer rats. Or 5 mg/kg
3x/wk produced thymic invol in adults and preweanlings.
produced thymic invol in adults and preweanlings. NK
suppressed in pups only at 10 mg/kg.
20 to 80 mg/kg TBTO in food to rats /6wks, dose response
NK activity suppressed in lung tissue.
MLR suppressed in adults at 20 mg/kg and at 10 mg/kg in
pups.
Adapted from Bozelka and Salvaggio 1985
* = prenatal exposure
-------
TABLE 13 (Cont.)
COMPOUND
HCB
Lead
Lindane
(b-HCH)
Mercury
CITATION
Renana and Rao 1992
Bamett el aL 1985*
Van Loveren el aL 1990
Bozelka and Salvaggio 1985
Buchmuller-Rouiller el aL 1989
Malviya el aL 1989
Cornacoff el aL 1988
WHO 1976
Contrino el aL 1988
Reardon and Lucas 1987
Blakley 1990
van Velsen el aL 1986
Mirtcheva el aL 1989
Rossert el aL 1988
Stiller- Winkler el aL 1988
Reardon and Lucas 1987
USPHS. ATSDR 1988
EFFECT
Immunosuppressive in prenatal mice.
150 rag/kg to 450 in food 6 weeks suppressed NK activity
dose response in rat lungs.
1 gm/d for 7d PbNOS increased susceptibility to Ascaridia
gallia.
Thymus weight loss
0.5 mg HgC12/kg bw s.c.3x/wk. Autoimmune response in
female rats.
100 u.g HgC12/100 g bw s.c.3x/wk. Autoimmune response in
male and female rats.
3 u.g Hg2 s.c. in murine hind foot pad stimulated
T-cell-dependent enlargement of the popliteal lymph node
(PLN).
Induces cytotoxic T-cells and interferon production in mice.
Induces glomerulonephritis in rats.
-------
TABLE 14
IMMUNOSUPPRESSIVE EFFECTS OF POLYCHLORINATED BIPHENYLS
COMPOUND
PCBs
PCBs
PCBs
PCBs
PCBs
PCBs
PCBs
PCBs
PCBs
PCBs
PCBs
PCBs
SPECIES
Monkey
Monkey
Monkey
Mouse
Rat
Monkey
Guinea
Pig
Rat
Chick
Quail
Guinea
Pig
Mallard
duck
EFFECTS
Increased natural killer cell
activity, interferon levels, and
thymosin alpha- 1 levels
Decreased IgM and IgG response
Reduced antibody levels
Inhibited splenic plaque-forming
cell response
Reduced activity of natural killer
cells, reduced thymus weight
Lowered antibody response
Reduced leukocytes and
lymphocytes, induced thymic
atrophy
Suppressed T-cell response
Inhibits lymphoid development in
the bursa of Fabricus
Immunosuppressive response
Immunosuppressive response
Immunosuppressive response
REFERENCE
Tryphonas et aL 1991a
Tryphonas el aL 1991b
Tryphonas el aL 1989
Howie el aL 1990
Smialowicz el aL 1989
Colborn 1989
Colbora 1989
Kerkvliet & Baecher-
Steppan 1988
Nikolaidis el aL 1988
Dieter 1974
Vos & De Roy 1972
Friend & Trainer 1970
49
-------
The mechanism of thymic involution in mammals is poorly understood. In mice, TCDD
severely impairs fetal liver and neonatal bone marrow prothymocyte activity, thereby disrupting
the seeding of the thymus with prothymocytes (Fine el aL 1989, 1990a, b). TCDD administered
to pregnant mice inhibits thymocyte maturation in embryos m utero (Blaylock cl aL 1992) and
decreases the number of thymic glucocorticoid receptors in both male and female rats in later life
(Csaba el aL 1991).
The effects of TCDD on mature immune cells are diverse. Although TCDD increases
natural killer cell (a type of T-cell) activity in the blood and spleen of mice, it decreases the
proliferative response of spleen lymphocytes (Funseth and Ilback 1992). TCDD acts by
impairing the function of helper T-celis, leading to an impairment of B-cell activation (Neubert
el aL 1990; Tomar and Kerkvliet 1991; Lundberg el aL 1991), and suppression of B lymphocyte
maturation and antibody synthesis (Clark fit aL 1991). This is accomplished by alterations in
tyrosine kinase activity that occurs within minutes of TCDD treatment (Clark el aL 1991).
However, House el aL (1990) noted a dose-dependent decrease in activity in both T-dependent
and T-independent antibody (IgM and IgG) forming cells.
In general, TCDD-induced immunosuppression requires induction of cytochrome
P4501A1 (Gasiewicz and Rucci 1991). However, certain aspects of immunosuppression may
operate through different mechanisms. Both T-helper cell and cytotoxic T lymphocyte activity
disruption may be independent of TCDD binding to the Ah receptor (Kerkvliet el aL 1990a, b).
Mercury exposure can impair immune system function by altering the activity and levels
of immune cells. Exposure via the placenta and milk impairs natural killer cell function in rats
(Ilback el aL 1991). Immunosuppressive effects, including a 22 percent decrease in thymic
weight and 50 percent reduction in thymic cells, occurred following 12 weeks of 3.9 ng/g oral
dosing in mice (Ilback 1991). Mouse splenic T lymphocytes were activated to display
cytotoxicity and produce interferon at lOu, of Hg+* (Reardon and Lucas 1987).
Mercury induces a significant autoimmune disease effecting the kidneys. Mercury
exposure leads to production of antibodies to renal basement membranes, resulting in
glomerulopathy (Bellon el aL 1982; Bernaudin el aL 1981; Andres 1984; Knoflach el aL 1986;
Fukatsu el aL 1987; Guery el aL 1990; Pelletier el aL 1990; Pusey el aL 1990; Hultman and
Enestrom 1992). Mercuric mercury, but not methylmercury, induces synthesis of metallothionein
by the kidney cells only (Amdur el aL 1991).
2.2.5.4 Metabolic Impairment
Metabolic changes as the result of exposure to chemical contaminants have been
documented in the mixed function oxidase (MFO) enzyme system of invertebrates, fishes, birds,
and mammals (Table 15). Functionally, this system acts to metabolize steroid hormones and
50
-------
xenobiotics for excretion. MFO enzymes such as aryl hydrocarbon hydroxylase (AHH) and
ethoxyresorufin-O-deethylase (EROD) are found in the liver, kidney, intestines, and most body
tissues. They respond to the presence of chemicals such as PCBs, PAHs, dioxins, and
organochlorine pesticides. Although the elevation of MFO enzymes is not necessarily an
indication of toxicity, it is an indicator of the presence of these particular substances and can be
used as a measure or biomarker of toxic exposure (Rattner el aL 1989). The biological responses
to AHH and EROD activity have been associated linearly with a number of toxic responses
including body weight loss ("wasting") and thymic atrophy in rats, cleft palate in mice, and mild
to severe porphyria, depending upon the species of animal exposed (Mason el aL 1985; Mason
el aL 1986; Mason el aL 1987). In some instances, the metabolic product of the enzyme activity
is more toxic than the original compound. Field investigators have used MFOs as measures of
xenobiotic exposure and in several instances have shown an association between elevated enzyme
activity and an adverse effect (Table 16).
Fish and Wildlife Studies
It has been suggested that MFO activity in a species is inversely related to the
accumulation of an enzyme inducing xenobiotic in a species, i.e., MFO activity level may
contribute to the amount of xenobiotic accumulated. Fish and fish-eating birds exhibit the lowest
MFO activity; other birds are intermediate; and mammals have the highest activity (Rattner el
aL 1989).
fish
The National Benthic Surveillance Project (Varanasi 1989) reported metabolic disorders
in fish from contaminated areas. A suite of metabolic bioindicators of contaminant exposure was
field tested in three species of Puget Sound fish: English sole (Parophrys vetulus); starry flounder
(Platichthys stellatus); and rock sole (Lepidopsetta bilineata), from five sites over a
contamination gradient. Comparisons of the concentrations of 24 aromatic hydrocarbons and
PCBs in sediment, fish liver PCB concentration, and fish bile fluorescent aromatic compounds
(FACs) (a bioindicator of contamination and metabolite accumulation in fish bile) were made on
seasonally-controlled samples. Although the results showed variation in response between tests,
all were sensitive enough to differentiate the levels of contamination. The National
Oceanographic and Atmospheric Administration (NOAA) also demonstrated a statistical link
between aromatic contaminants and other metabolic effects such as induction of the MFO
cytochrome P450 enzyme system in field and laboratory studies of the following fish: Atlantic
croaker, black croaker, California halibut (Paralichthys californicus), Chinook salmon
(Oncorhynchus tshawytscha), Coho salmon (Onchorhynchus Idsutch), Dolly Varden (Salvelinus
malma), English sole, flathead sole, hardhead catfish, hornyhead turbot, Pacific halibut
(Hippoglossus sp.), rock sole, starry flounder, white croaker (Genyonemus lineatus) and winter
flounder (Pseudopleuronectes americanus) (NOAA 1991).
MFO activity in lake trout (Salvelinus namaycush) and white suckers (Catastomus
commersoni) from Lakes Ontario and Michigan was higher when compared with activity in fish
51
-------
from Lakes Superior, Erie, and Huron (Hodson el aL 1989). MFO activity in Lake Michigan lake
trout embryos as a result of parental exposure was 3.5 to 6.5 times higher than hi embryos from
hatchery stock. MFO activity abated in the embryos after several months hi clean water (Binder
and Lech 1984).
Wildlife.
"Wasting" and egg shell thinning in colonial nesting birds were described among the
earliest reports of wildlife damage in the Great Lakes (Gilbertson 1975). Ellenton and coworkers
(1985) were the first to use enzyme induction as a measure of exposure as well as toxicity in
field research (Table 17). Exposure to organic contaminants has been associated with MFO
activity hi birds and reptiles as well as fish.
Custer and Peterson (1991) studied black-crowned night-heron (Nycticorax) MFO activity
to determine its applicability for use as indicators of U.S. estuarine contamination. Enzyme
activity and pollutant load in black-crowned night-heron chicks hi Chincoteague National
Wildlife Refuge were compared with chicks from more polluted sites in Green Bay, Wisconsin
and San Francisco Bay, California. In comparison to the Chincoteague reference site, San
Francisco Bay chicks displayed significantly greater AHH activity.
Porphyria, a condition wherein heme biosynthesis is altered, results hi the accumulation
in the liver of porphyrins, precursors to hemoglobin. HCB, PCBs, and dioxins induce the
accumulation of highly carboxylated porphyrins (HCPs) and are measurable hi liver tissue and
the blood. Their presence is used specifically as an indicator of exposure to PCBs, HCB, and
TCDD (Marks 1985). The porphyrins are toxic and are components of the suite of lesions for
diagnostic chick edema disease (Gilbertson 1992). The levels serve as distinct measures of
change hi the presence of the above organochlorine chemicals. The Canadian Wildlife Service
has plotted the variation hi highly carboxylated porphyrins in herring gulls from various locations
around the Great Lakes (See Figure 5).
HCB caused porphyria cutanea tarda (PCT) in children, one year of age or less, whose
mothers consumed HCB-treated wheat hi an incident in Turkey during a famine (Jones and
Chelsky 1986). All children exposed in utero expired within two years after birth.
The U.S. National Human Adipose Tissue Survey (Murphy ei aL 1983) and a nationwide
breast milk study in Canada (Davies and Mes 1987) found HCB 100 percent of the time in fat
and breast milk, respectively. A recent report indicates that HCB over the ten year period
between 1975 and 1985 remained constant or possibly increased hi human adipose tissue (OWRS
1986). Regular fish eaters hold higher concentrations of HCB than lacto-vegetarians and mixed
dieters (Noren 1983).
52
-------
TABLE 15
ENZYME TEQ IN GREAT LAKES ANIMALS
Dioxin enzyme induction toxicity screening (TCDD equivalents) and specific dioxin and PCB congeners for which dose-response
associations have been made with morbidity and mortality in wildlife populations.
Biologic Marker
TCDD equivalents
3,4,5,3',4'-penta PCB
3,4,3',4'-tetra PCB
2,3,7,8-TCDD
TCDD equivalents
TCDD equivalents
3,4,5,3',4'-penta PCB
3,4,3',4'-tetra PCB
2,3,7,8-TCDD
TCDD equivalents
. 3,4,3',4'-tetra PCB
TCDD equivalents
TCDD equivalents
Wildlife Species
Forster's tern
Caspian tem
DC Cormorant
Lake trout
Coho salmon
Herring gull
Mortality and Morbidity Endpolnts
embryonic mortality
deformities
embryonic mortality
deformities
embryonic mortality
deformities
egg mortality
hatchability
embryonic mortality
embryonic mortality
deformities
Citations
Kubiak el aL 1989
Ludwig and Giesey 1990
Giesey el aL 1991
Ludwig and Giesey 1990
Giesey el aL 1990
Tillit el aL 1992
Mac and Edsall 1989
Ludwig and Giesey 1990
Ludwig and Giesey 1990
53
-------
TABLE 16
REVIEW OF MECHANISM OF ACTION OF COMPOUNDS OF CONCERN
ACTIVITY AND CITATION
COMPOUND
2,3,7,8-TCDD
B[a]P
Chlordane
DDE
Dieldrin
HCB
Lead
ENZYME INDUCERS
Silbergeld and Mattison 1987
Bradlaw and Casterline 1979
Traber et aL 1988 (intest.),
Haake el aL 1987
Bulger and Kupfer 1983,
Haake el aL 1987
Haake et aL 1987
Gutkina and Mishin 1986,
Stewart and Smith 1986,
Haake el aL 1987
*
INHIBITORS OF GAP s
JUNCTIONAL COMMUNICATION
•
Zhong-Xiang et aL 1986,
Warngard el aL 1988,
Trosko and Chang (in press),
Klaunig and Ruch 1987a, b,
Ruch el aL 1987 (DDT)
Zhong-Xiang et aL 1986
DISRUPTION OF
ENDOCRINE CONTROL
Umbreit and Gallo 1988,
Silbergeld and Mattison
1987,
Gallo 1988,
Romkes and Safe 1898
Cranmer et aL 1984,
Welsh el aL 1971
Fry et aL 1987,
Rattner et aL 1984,
Bulger and Kupfer 1983,
Fry and Toone 1981,
Lundberg 1973
Haake et aL 1987
Haake et aL 1987,
Elissalde and Clark 1979
Rodamilans et aL 1988,
USPHS. ATSDR 1988
-------
Laboratory And Mechanistic Studies
This section will deal with certain effects on systemic, cellular, and biochemical
metabolism. Xenobiotics have an enormous effect on the body by their induction of metabolic
enzyme systems. These enzymes regulate the metabolism of many endogenous chemicals, such
as hormones, and foreign contaminants as well.
Systemic metabolic depression leading to slow starvation and eventual death is referred
to as the wasting syndrome. The mechanistic basis of the wasting syndrome has proven to be
particularly elusive. There are several different mechanisms by which the anorexia (loss of
appetite) and hypophagia (decrease in food intake) of the wasting syndrome may occur. These
include enzymatic induction of the mixed-function oxidase (MFO) system, neurological changes,
and disruption of several different endocrine hormones, receptors, and feedback mechanisms. It
is likely that the wasting syndrome is a manifestation of multiple biological effects. Refer to
Table 17 for a summary of the different mechanisms implicated in the wasting syndrome.
The body has natural defenses to eliminate foreign compounds from its system. Many
substances that are water soluble are rapidly eliminated by the kidneys and tend not to
bioaccumulate. Alternately, organic compounds are less water-soluble, and are far more difficult
to excrete. Organic xenobiotics are therefore oxidized to form water-soluble metabolites that can
be further conjugated and excreted in the urine or bile (Lech el aL 1982; Payne el aL 1987). The
major means of xenobiotic oxidation are accomplished through a complex metabolic pathway
referred to as the mixed-function oxidase system.
The mixed-function oxidase system, or MFO, is located in the microsomal portions of
various tissues, especially of the liver. It is characterized as comprising an electron transport
system with cytochrome P450, requiring NADPH (or NADH) as a cofactor, and being capable
of oxidizing many different kinds of substrates (i.e., substrate nonspecificity). Cytochrome P450
is the component of the MFO system that actually binds to both oxygen and substrate molecules.
Other enzymes, such as NADPH-cytochrome-c-reductase (a flavoprotein) mediate the transport
of electrons from NADPH to cytochrome P450.
Cytochrome P450 consists of a family of hemoproteins called monooxygenases. The
entire system of monooxygenases collectively forms the MFO system. In humans there are over
30 different cytochrome P450s identified (Guengerich 1992). Many monooxygenases are capable
of oxidizing different substrates (Guengerich 1991). This enables the cytochrome system to
oxidize many different natural substances, as well as xenobiotics. Natural substrates in the body
include steroid hormones, prostaglandins, fatty acids, leukotrienes, biogenic amines, pheromones
and plant metabolites (Nebert and Gonzalez 1987).
The MFO system is the bodies first line of defense against xenobiotics (Payne el aL
1987), including many drugs, chemical carcinogens, mutagens, and environmental contaminants
(Nebert and Gonzalez 1987). The induction of monooxygenases is relatively non-specific. A
single xenobiotic can induce the production of many members of the cytochrome system. For
56
-------
TABLE 16 (Cent.)
ACTIVITY AND CITATION
COMPOUND
ENZYME INDUCERS
INHIBITORS OF GAP
JUNCTIONAL COMMUNICATION
DISRUPTION OF
ENDOCRINE CONTROL
Lindane [g-HCH]
Zielmaker and Yamasaki 1986,
Ruch el aL 1987 (g-HCH),
Trosko and Chang (in press)
Uphouse 1987,
Van Velsen el aL 1986,
Van Giersbergen el aL
1984
Lindane [b-HCH]
Schroter el aL 1987,
Van Velsen el aL 1986
Van Velsen el aL 1986,
Van Giersbergen el aL
1985
Mercury
Veltman and Maines
1986,
USPHS-ATSDR 1988, p.
57
Mirex
WHO 1984
Carlson el aL 1985,
Rosenbaum and Charles 1986,
Trosko and Chang in press
PCBs
Safe 1984,
Mason el aL 1986, 1987,
1988,
Traber el aL 1988 (intest.)
Tsushimoto el aL 1983,
Ruch el aL 1987 (Aroclor 1254),
Trosko and Chang (in press)
Dieringer el aL 1979,
Biessmann 1982
Toxaphene
Haake el aL 1987,
WHO (Camphechlor) 1984,
Chu el aL 1988
Trosko and Chang (in press)
Mohammed el aL 1985,
WHO (Camphechlor)
1984
-------
example, seven different Cytochrome P450s may be induced by barbiturates (Guengerich 1992).
This makes the MFO system capable of responding to a wide variety of xenobiotics. Further,
once the MFO system is induced by one xenobiotic, it is capable of rapidly responding to others.
This also makes MFO induction one of the most sensitive physiological indicators of
environmental pollution (Payne el aL 1987; Narbonne 1991; van der Oost el aL 1991; Pesonen
el aL 1992). MFO systems are wide-spread among species, although there is considerable
variability in specific enzymes (Nebert el aL 1981).
The mechanism of MFO induction is best understood for dioxins (Figure 6). For TCDD
to produce an effect, it must bind to the aromatic hydrocarbon (Ah) receptor, forming the
inducer-receptor complex that is transported to the nucleus by the Ah receptor nuclear
translocator protein (ami) (Reyes el aL 1992). The inducer-receptor complex subsequently
interacts with one or more of the Ah-responsive elements (AhREs) located upstream from the
transcriptional initiation site (Carrier el aL. 1992). Transcription of a gene such as CYP1A1
(cytochrome P4501A1) requires phosphorylation by protein kinase C in order to form a
transcriptional complex (Carrier el aL 1992).
CYP1A1 and its associated enzyme product, the aryl hydrocarbon hydroxylase (AHH)
assist in detoxification of polycyclic aromatic hydrocarbons (Safe 1986; Landers and Bunce
1991). The CYP1A1 gene exhibits differences in induction response between males and females
(Jones el aL 1991). Microsomal enzyme activity may be markedly increased in females, but
limited hi males. Vitamin C (ascorbic acid) reduces the microsomal aryl hydrocarbon
hydroxylase (AHH) activity induced by TCDD hi mice (Kiyohara el aL 1991). Alternately, PCBs
increase cellular levels of ascorbic acid (Nagaoka el aL 1991).
PCBs induce, in hepatic microsomes in vivo, a variety of different forms of the
cytochrome P450 enzyme systems involved in the metabolism of xenobiotics (Borlakoglu el aL
1990). This includes increases in cellular levels of AHH (Nagaoka el aL 1991). PCBs covalently
bind to DNA following metabolic activation, although the more highly chlorinated congeners are
poorly metabolized in vivo and do not readily form covalent adducts (Safe 1989). A linear
association exists between PCB dose and cytosolic protein binding; between protein binding and
enzyme induction; and between enzyme induction, immune suppression, teratogenicity, and
wasting (Safe 1984; Safe el aL 1985; Mason el aL 1986; Mason el aL 1987).
57
-------
TABLE 17
MIXED FUNCTION OXYGENASE RESPONSES DOCUMENTED IN FREE-RANGING WILDLIFE
MFO Response
Species Age Sex Site Comparison or Tissue Residue or Type Change Refer-
Control Potential ence
Exposure
Herring gull
Herring gull
Forster's
tern
Herring gull
i
Black-
crowned
night-heron
American
robin
•
20- and
25-day
old
embryo
25-day
embryo
1 -day-
old
hatchling
21 and
25 day
embryo
2, 7, 1,
and 21
day-old
nestling
Pipping
embryo
Adult
—_
__.
__
Male
and
Female
—
—
Great Lakes
Great Lakes
•
Great Lakes
Newfoundland
San Francisco Bay
Pine plantations in
Wisconsin
Association between
MFOs and residues;
unpolluted control site
Association between
MFOs and residues;
unpolluted control site
Unpolluted control site
Association between
MFOs and residues
Association between
MFO's and residues;
captive control
Unpolluted control site
Pentachlorobenzene
TCDD
DDE, Mirex,
Hexachlorobenzene
,PCBs
PCBs, TCDDs,
Polychlorinated
dibenzo-p-dioxins
DDE, Dieldrin,
Heptachlor
epoxide,
Oxychlordane,
Hexachlorobenzene
and PCBs
Organochlorines,
PCBs
TCDD,
Polychlorinated
dibenzo-p-dioxins
AHH
EROD
APDM
AmH
AHH
AHH
APDM
EROD
Cytochrome
P-450
AHH
AHH
EROD
+
+
-
-
+
0
0
0
+
0
+
+
Ellenton
•elaL
1985
Boersma
eiaL
1986
Hoffman
elaL
1987
Peakall
elaL
1986
Hoffman
fitaL
1986
Martin
elaL
1987
-------
TABLE 17 (Cont.)
MFO Response
Species Age Sex Site Comparison or Tissue Residue or Type Change Refer-
Control Potential ence
Exposure
Double-
crested
connorant
Razorbill
and puffin
Pigeon
Black-
necked
grebe
Black-
headed gull
Cotton rat
Adult
—
Adult
Adult
Male
and
Female
—
__
Male
Great Lakes
Saltee Islands,
Ireland
Isle of May and
Outer Hebrides,
Scotland
Lucknow, India
Marano, Italy
Central Italy
Texas
Across a geographic
pollution gradient
Association between
MFOs and residues
Reared in captivity
Various intervals of
residence at polluted site
Association between
MFOs and residues;
dump versus lagoon
Unpolluted control site
PCBs
PCBs
DDE, DDT,
Hexachlorocyclohe
xane, Lindane
DDE,
Hcxachlorobenzene
*
PCBs
PCBs
Arsenical
herbicides,
Dieldrin, Petroleum
hydrocarbons,
PCBs
Aldrin
epoxidase
Hyrdoxyla-
tion of
dieldrin
analogue
AHH
Aldrin
epoxidase
EROD
Aldrin
epoxidase
EROD
AnH
Cytochrome
P-450
+
0
0
+
+
+
+
+
0
Tillet el
aL1992
Knight
and
Walker
1982
Kaphalia
eiaL
1981,
Husain el
8L1981
Fossi
eiaL
1986
Fossi
eiaL
1986
Rattner el
aL 1986,
Rattner el
aL1987
-------
Mercury is a potent, nonspecific enzyme poison. It produces its effects by releasing
mercuric ions, which readily form covalent bonds with sulfhydryl groups (Winek el aL 1981).
This results in the inhibition of metabolic enzymes, denaturation of proteins, and disruption of
cell membranes (Bryson 1989; Chetty el aL 1990; Gill 1990; Boadi el aL 1991; Dieter el aL
1992; Wigfield and Eatock 1992; Anner and Moosmayer 1992; Suresh el aL 1992). However,
methylmercury does induce AHH activities (Boadi el aL 1991, 1992).
Metabolism of xenobiotics is normally thought of a "detoxification," but this is not always
the case. Sometimes, in the body's attempt to rid itself of foreign materials, it actually creates
reactive intermediates that are more toxic than the original compound (Anders 1985; Thakker el
aL 1985; Nebert and Gonzalez 1987; Butler el aL 1989; Aoyama el aL 1990; Guengerich 1992).
This type of transformation is referred to as "activation". P450 cytochromes are involved in the
metabolic activation of polycyclic aromatic carcinogens (Fujii-Kuriyama el aL 1990).
Further, by inducing the MFO system, xenobiotics stimulate changes in enzymes
regulating other body functions. Associated with the wasting syndrome are changes in
carbohydrate homeostasis. Correlated with the reduction in feeding is a decrease in formation
of the essential blood sugar glucose (gluconeogenesis) by the livers of rats exposed to TCDD.
Both appetite lose and reduction of hepatic enzyme activity occurred in the same dose ranges,
suggesting a possible cause and effect relationship (Weber el aL 1987,1991). In birds, TCDD-
induced wasting is associated with impaired carbohydrate production (Lentnek el aL 1991). In
human cells, TCDD completely inhibited the conversion of glucose to lactate (Narasimhan el aL
1991).
Changes in regulatory enzymes of the MFO system affect other systems as well.
Particularly important are changes in sex steroid levels that influence reproductive cycles,
behavior, and fertility. These effects of xenobiotics on behavior and reproduction will be
discussed under the appropriate section.
60
-------
FIGURES
PORPHYRIN LEVELS IN LIVERS OF GREAT LAKES HERRING GULLS IN 1985
Figure 3
topftyrin l«v«ta In Uvws of Oraot LokM hwitng guto In 1985. Madia* tov«te o< higMy
eaooxyidcd poreftyiin i*v«a in Uv«a con*ct«a (ram s*v«n G««at u«« cotootM «•
**!*«M«
-------
FIGURE 6
MECHANISM OF DIOXIN-Ah RECEPTOR ACTION
TCDD
Cytosol
TCDD
Binding to
Receptor
Transformation
Transportation
byamt
Ah
Receptor
P-450
Activated
Receptor
Translation
, Transcription
Gene
and other
Proteins
I
mRNA
Multiple Biological
Responces
DMA
Source: Modified from Landers and Bunc*. 1991
Proposed mechanism of dioxin action through the Ah-receptor. TCDD enters the cell where it is bound by the Ah-receptor (aromatic
hydrocarbon) molecule. TTie TCDD and its bound receptor are transformed into an activated complex, which is transported into the
nucleus by arm (Ah-receptor nuclear translocator protein). The activated complex binds to the AhRE (Ah-rcsponsivc elements)
enhancing transcription of structural genes into mRNA (messenger RNA). The mRNA is translated into several cytochrome P-450
enzymes and other proteins, resulting in an array of biological responses.
-------
TABLE 18
MECHANISMS IMPLICATED IN THE WASTING SYNDROME
Target Organ
Liver
Brain
Thyroid
Adrenal Gland
Pancreas
Mechanism
Mixed-function Oxidase System
Neurotransmitters
Thyroxine & Triiodothyronine
Corticosterone
Insulin & Glucagon
Process Affected
Carbohydrate Metabolism
Feeding Behavior
Cellular Metabolism
Brown Fat Metabolism
Gluconeogensis
Blood Glucose Levels
2.2.5.5 Nervous System and Behavioral Impairment
Wildlife Studies
Overt and subtle behavioral changes have been identified in wildlife and human
populations who consumed contaminated fish. Wildlife populations exhibited changes in sexual
and nesting behavior (Burger 1990; Conover 1984; Conover and Hunt 1984a, 1984b; Kovacs and
Ryder 1981; Kubiak el aL 1989; Fox and Weseloh 1987; Fry el aL 1989; Fry and Toone 1981;
Nisbet and Drury 1984; Shugart el aL 1988). Diamond (1989) points out that these changes in
sexual behavior were not reported before 1950 in aquatic birds. The onset of these changes
coincides with the first reports on eggshell thinning and gross mortality in wildlife populations
around the Great Lakes (Colbom 1988) and supports the hypothesis that post World War n
chemical production has an influence on ecosystem health (Colbom 1991) (Figure 7).
Populations of Great Lakes herring gulls, Forster's terns, and ring-billed gulls suffering
reduced reproductive success also exhibited behavioral changes such as female-female pairings,
aberrant incubation activities, and nest abandonment (Shugart el aL 1988; Fox and Weseloh
1987). Female-female pairings of herring gulls resulted in supernormal clutches, 4-8 eggs per
nest rather than 3 eggs (Fox and Weseloh 1987; Peakall and Fox 1987). Although egg-laying
capacity was not impaired, only 10 to 30 percent of the eggs were fertile (Shugart el aL 1988).
63
-------
Nest abandonment was observed and hatching success was reduced in Green Bay (26
percent) versus inland (88 percent) Forster's tern colonies (Hoffman el aL 1987; Kubiak el aL
1989). Fox el aL (1978) found a positive correlation between abandonment (time unattended)
of Lake Ontario herring gull nests and the level of contaminants in the eggs. Follow-up egg-
swapping field studies for both the herring gull and Forster's tern determined that extrinsic
parental behavior contributed to the intrinsic factors also affecting reproduction (Peakall el aL
1980; Kubiak el aL 1989). For a description of the Forster's tern study see Section 2.2.5.3. In
a herring gull study on Lake Ontario in the early 1970s, Peakall and coworkers (1980) found that
contamination levels of the colonies determined hatchability.
Supernormal clutches were also observed in the ring-billed, California, and western gulls
of Oregon and Washington (Conover 1984; Conover and Hunt 1984a, b). Increase in female-
female pairing correlated with the reduction in numbers of male birds during the breeding season.
A frequency of double-nests and/or supernormal clutches (0.0005-0.01 percent) in New England
herring gulls was compared with Great Lakes and West Coast observations (0.3 percent) (Nisbet
and Drury 1984).
The New England gulls held little or no detectable DDT. Hunt and coworkers (1980)
reported an incidence of 14 percent in female-female pairing among western gulls on Santa
Barbara Island, California. Using museum specimens, Conover and Hunt (1984a) sexed post-
1950 and pre-1940 western gulls and found a significantly lower male to female ratio in the
post-1950 birds. Fry and Toone (1981) demonstrated that feminization (abnormal growth of
oviducts and ovarian tissue) of male embryos occurred with exposure of wild adults in the field
to DDT. The reduction in breeding male birds leading to female-female pairing and supernormal
clutches was hypothesized to be from increased male mortality or feminization of male birds
from contaminant exposure (Conover and Hunt 1984b; Nisbet and Drury 1984; Fry and Toone
1981).
Other behavioral change from DDT metabolites and DDT analogs was demonstrated in
experiments on the American kestrel (Falco sparverius) with in ovo exposure to p,p'-dicofol
(registered name Kelthane), a structural analog of DDT and DDE (Fry el aL 1989). First and
second generation studies resulted in the following: testicular feminization of first generation
males from Kelthane, dicofol, and DDE exposure; and a dose-response reduction of male
aggressive behavior and infertility from Kelthane.
Adult rats fed a 30 percent diet of salmon from the Salmon River, a tributary to Lake
Ontario, developed an aversion to stress after 20 days (Daly 1989). All the rats fed Lake Ontario
salmon were hyper-reactive to stressful events such as reductions in food rewards, mild shocks,
and novel environments compared with rats fed Pacific salmon or no salmon. The same effects
were seen after a 10 percent diet fed for 60 days (Daly 1991). In a later study, female rats were
fed Lake Ontario, Pacific, or no salmon from the day they were placed with males until their rat
pups were 7 days old. Their pups continued to nurse until 21 days old and were never fed Lake
Ontario fish. Nonetheless, all pups from dams fed Lake Ontario fish exhibited hyper-reactivity
to stressful events when tested as juveniles and as adults (Daly 1992b). Total PCBs and mirex
64
-------
FIGURE?
EFFECTS REPORTED IN GREAT LAKES WILDLIFE SINCE WORLD WAR H
e^
_
Effects n
Adapted
Bald Eagle
Beluga Whale
Black-Crowned NH
Caspian Tern
Chinook/Coho Salmon
Common Tern
D.C. Cormorant
Forster's Tern
Herring Gull
Lake Trout
Mink
Osprey
Otter
Ring-Billed Gull
Snapping Turtle
X
X
X
X
N/A
X
X.
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
N/A
X
N/A
X
X
N/A
N/A
X
N/A
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
N/A
X
X
X
X
X
X
X
X
Observed effects that have been reported in the literature.
sported in Great Lakes wildlife since World War II in populations dependent upon fish from the lakes.
from: Colborn (1991)
-------
were the only contaminants quantified in the fish and the brains of the rats in the studies
(Hertzler 1990). Both contaminants were significantly higher in the Lake Ontario fish and rats
on the Lake Ontario fish diet compared with the Pacific Ocean fish- and mash-fed rats. In
concurrent studies, researchers demonstrated an inverse association between tissue dopamine
production and several non-dioxin-like PCB congeners (2,4,4'; 2,4,2',4'; 2,4,2',4',6') found hi the
fish Daly fed her rats (Seegal el aL 1985; Bush el aL 1990; Shain el aL 1990; Seegal 1992a, b).
The children of women who consumed Lake Michigan fish two to three times a month
exhibited subtle changes in cognitive processing and altered activity levels (Jacobson el aL 1985;
Rogan el aL 1988; Swain 1988; Jacobson el aL 1989; Winneke el aL 1989; Jacobson el aL 1990;
Tilson el aL 1990; Jacobson el aL 1992). Children accidentally exposed in utero to cooking oil
contaminated with PCBs and dibenzofurans exhibited similar neurological decrements (Rogan el
aL 1988). Similar psychomotor events were documented in a North Carolina cohort of infants
whose mother's milk delivered equivalent levels of PCB as those determined in the Lake
Michigan mother's milk (Rogan el aL 1986). In each study neurological events were observed
at the same level of PCB in breast milk. However, the neurological changes appeared not to
persist in the North Carolina cohort as they did in the Lake Michigan cohort. Different
instruments were used for testing in the two studies.
An association was found between the activity level in four-year old breast—fed children
and concentrations of PCBs in the mothers milk (Jacobson el aL 1992). The children were
exposed to elevated levels of PCBs as the result of their mothers' Lake Michigan fish
consumption or their mothers' having consumed PCB-spiked farm products via contaminated
silage. Hypotonicity and hyporeflexivity were increased in those children who nursed for more
than a year and whose mothers' milk held the highest concentrations of PCB. Mothers' milk with
PCB levels exceeding 1000 ppb contributed 0.19 ± 0.03 ppb per week to the offspring's serum
at age 4. Mean serum concentration at 4 years was 5.1 ± 3.9 ng/ml in children who breast fed
for 6 months, 1.2 ± 1.6 ng/ml for less than 6 months, and 0.3 ± 0.7 ng/ml for those who did not
breast feed. In both cohorts, growth retardation as the result of in utero exposure persisted in
a dose-dependent manner through age four and was observed, along with the neurotoxicological
effects. Reduction in activity was also related to the youngsters' PCB body burden. The effects
were more pronounced hi females than males. Seventeen of the breast-fed children, all from
mothers' with high PCB milk concentrations, refused psychological testing. This finding is
consistent with the rat studies cited above (Daly 1992a).
Using the results of laboratory animal studies and the Jacobsons' studies, Tilson el aL
(1990) determined that, neurotoxicolologically, humans are four orders of magnitude more
sensitive to PCBs than rodents. In their analysis, they found that contemporary levels of PCBs
transferred to human offspring in utero were associated with "...hypotonia, hyporeflexia at birth,
delay in psychomotor development at 6 and 12 months, and poorer visual recognition memory
at 7 months" (p. 239). The above effects are not visible and would ordinarily go undetected.
In this case, skilled psychologists, unaware of the exposure history of the child, detected the
effects in the children of women who ate Lake Michigan fish. These effects were found in the
children of women who represented the upper 95 percent in a normal population based on PCB
exposure.
66
-------
Laboratory And Mechanistic Studies
There are many different types of behavioral impairment brought about by xenobiotic
contaminants. Some affect reproductive behaviors, ranging from inappropriate courting and
mating behaviors to miscare of eggs or young. Others involve the anorexia and hypophagia
associated with the wasting syndrome. It is apparent that xenobiotic contaminants operate
through a variety of neurologic mechanisms that ultimately lead to behavioral impairment.
The treatment of animals with xenobiotics brings about many of the behavioral
abnormalities seen hi wildlife from polluted areas. Feeding ring doves mixtures of DDE, PCB,
and mirex produced behavioral abnormalities similar to those observed in Lake Ontario herring
gulls (i.e., abnormal incubation behavior). These effects were dose-related to decreases in
circulating androgens in males, estrogens and progesterone hi females, and thyroxine hi both
sexes. Prolactin (which influences behavior in many vertebrates) was also altered in some
individuals (McArthur el aL 1983).
Many behavioral effects are not due simply to changes in the endocrine system, but to
direct effects of xenobiotics on the brain. In pigeons, 1 percent of PCBs administered was found
in the brain within 120 hours of treatment (Borlakoglu el aL 1991). PCBs have been shown to
accumulate in the brains of cod and trout (Ingebrightsen el aL 1990) and TCDD in the brain of
cod (Ingebrightsen el aL 1991). Administration of TCDD directly to the intracerebroventricular
fluid in rats produces significantly stronger reactions than peripheral administration, suggesting
that the central nervous system plays an important role hi TCDD toxicity (Pohjanvirta el aL
1989).
One of the most obvious effects of xenobiotics on wildlife is the wasting syndrome.
TCDD treatment of rats leads to a decrease in food intake (hypophagia) and aversion to eating
energy-providing foods. The neurological bases of altered satiety levels have been difficult to
deduce. Studies have linked TCDD-induced wasting in rats with increased levels of serotonin
(a neurotransmitter), or its precursor, tryptophan, in the brain (Rozman el aL 1991). However,
TCDD can cause wasting even if serotonin levels are artificially reduced (Stahl el aL 1991),
suggesting that factors other than serotonin are involved. Stahl and Rozman (1990) concluded
that the effect of TCDD does not involve the brain, but rather a peripheral appetite suppressive
(feedback) mechanism outside the central nervous system. Pohjanvirta and Tuomisto (1990a, b)
suggest that hypersensitivity of the central nervous system to peripheral satiety signals coupled
with hyporesponsiveness to metabolic energy deficit cues are involved in the wasting syndrome
mechanism.
Dopaminergic neurons of the brain are particularly sensitive to environmental and
pharmacological agents (Seegal el aL 1991a). The neurologic effect of PCBs and TCDD is
correlated with decreased levels of the neurotransmitter dopamine (Russell el aL 1988; Seegal
el aL 1991b). However, in rats exposed to 50 |ig/kg TCDD, only slight changes in dopamine
and several other aminergic neurotransmitters were noted from 4 to 76 hours following exposure.
Although TCDD causes changes in brain aminergic neurotransmitter systems, the changes were
67
-------
minor and it is unlikely that aminergic systems are solely responsible for TCDD-induced
hypophagia (Tuomisto el aL 1990).
The degree of PCB chlorination determines if dopaminergic functions will be altered in
the peripheral or central nervous systems (Seegal el aL 1988). Following exposure to Aroclor
1016, dopamine concentrations were significantly reduced in the brain of monkeys. Only three
PCB congeners (2,4,4'; 2,4,2',4'; and 2,5,2',5') were subsequently found in the brain. These
congeners were shown to reduce cellular dopamine concentrations in cells cultured in vitro,
whereas planar, dioxin-like congeners (3,4,4',4', and 3,4,5,3',4') did not (Seegal el aL 1990).
Studies in primates indicate that it is PCBs themselves, not their metabolites, that are responsible
for neurotoxic effects (Shain el aL 1991). These studies, both in vivo and in vitro, suggest that
PCBs may reduce dopamine concentrations through a novel mechanism and not through the Ah-
receptor complex responsible for both immunotoxic and hepatotoxic changes following exposure
to dioxin and dioxin-like PCBs (Seegal el aL 1990; Shain el aL 1991).
TCDD also may impair behavior and nervous system functions through disruption of
endorphins and their receptors. Endorphins are natural brain peptides exhibiting morphine-like
analgesic properties that may regulate behavior. TCDD causes perturbations in hypothalamic
beta-endorphin concentrations and brain mu opioid receptor numbers, which may contribute to
the mechanisms by which TCDD leads to decreased food intake and the wasting syndrome
(Bestervilt el aL 1991).
DDT and its analogs appear to alter behavior through both endocrine and neurological
mechanisms. The sexual (lordosis) behavior of adult female rats has been modified by single
dose exposure to DDT. Although both o,p'-DDT and p,p'-DDT decreased lordosis behavior,
they did so by different mechanisms. Whereas o,p'-DDT altered behavior by disrupting the
estrous cycle due to its estrogenic properties, p,p'-DDT had a major effect on the female's
proceptivity and receptivity without modifying her reproductive cycle (Uphouse and Williams
1989). Administration of p,p'-DDT decreased the level of the neurotransmitter serotonin within
hours of treatment (Uphouse el aL 1990).
DDT has a tremendous influence on development of the nervous system in embryos.
Neonatal exposure of mice to DDT caused changes in cholinergic receptors in the brain.
Subsequently, these same mice exhibited learning disorders as adults (Eriksson el aL 1990b). A
single oral dose of low-level DDT (1.4 mjimol/kg) to neonatal mice led to a permanent
hyperactive condition as adults (Eriksson el aL 1990a).
In every animal species studied, the nervous system is adversely effected by
methylmercury (WHO 1990). Further, methylmercury is one of the most potent neurotoxins
known (Pryor el aL 1983), and is readily transported across the blood-brain barrier (Aschner and
Aschner 1990; Kerper el aL 1992). Lesions are frequently observed in the granular layer of the
cerebellum (Herigstad el aL 1972; Falk el aL 1974; Chang 1977; Davies el aL 1977; Jacobs el
68
-------
aL 1977). In humans, the nervous system is the principal target of methylmercury exposure
(Who 1990; Amdur el aL 1991), with the fetus of exposed mothers being particularly susceptible
to deleterious effects (Cox el aL 1989). Damage to the brain is highly localized in the visual
cortex, granular layer of the cerebellum, and sulci (WHO 1991).
Prenatal exposure of offspring to doses that do not effect the mother produce abnormal
behavior in animals (Spyker el aL 1972; Bomhausen el aL 1980; Zimmer el aL 1980; Shimai and
Satoh 1985; Eisner el aL 1988). In monkeys exposed from birth to seven years of age, overt
behavioral effects were not manifested until they were 13 years old, demonstrating delayed
effects of mercury long after exposure (Rice 1990). Effects include hydrocephalus, decreased
cerebral cortex thickness, and increased hippocampus thickness (Kutscher el aL 1985).
Neurotransmitters and then* receptors in the brain are effected by mercury exposure
(Kobayashi el aL 1979, 1981; Concas el aL 1983; Atchison and Narahashi 1982; Quandt el aL
1982; Atchison 1986; Komulainen and Tuomisto 1987). Serotonin concentrations are increased
in rats following a single dose of 5.0 mg mercury/kg delivered as methylmercury on postnatal
day 2 (O'Kusky el aL 1988). Noradrenaline levels were increase significantly in the cerebellum
of rats 50 days following parturition when exposed to low doses (3.9 mg/kg in diet of dam)
during gestation and lactation (Lindstrom el aL 1991). The maturation of catecholamine
neurotransmitter systems in rats are adversely effected by early postnatal exposure (Bartolome
61 aL 1982).
The mechanism of mercury action in the brain is complex. In developing brains, some
effects are do to decreased motility of developing astrocytes (Choi and Lapham 1980), alterations
of cell membrane surface charge (Peckham and Choi 1986; Bondy and McKee 1991), disruption
of cell-cell recognition (Jacobs el aL 1986), and reduced myelination (Annau and Cuomo 1988).
Cell division is blocked during metaphase (Sager el aL 1982, 1983; Rodier el aL 1984; Slotkin
el aL 1985; Howard and Mottet 1986; Vogel el aL 1986) due to disruption of microtubules by
methylmercury (Imura el aL 1980; Sager el aL 1983; Miura and Imura 1987). Methylmercury
also disrupts levels of nerve growth factor in developing rat brains (Larkfors el aL 1991). Protein
synthesis also is impaired (Cheung and Verity 1985; Sarafian and Verity 1985, 1986; Thomas
and Syversen 1987). Male mice are more sensitive than females, which is consistent with
observations in humans (Sager el aL 1984; Choi el aL 1978).
There are a wide variety of neuronal and behavioral effects caused by xenobiotic
compounds (Table 19). These range from altering neurotransmitters and enzyme activities,
disordering cell membranes, impairing ion channels through membranes, and disrupting cellular
cytoskeletal elements. It is clear that we do not fully understand the mechanism of action of any
xenobiotic on the nervous system. A single xenobiotic may have many different effects, which
are brought about through, multiple mechanisms.
69
-------
TABLE 19
BEHAVIOR AND NEUROLOGIC EFFECTS OF XENOBIOTICS
COMPOUND
DDT
DDT
DDT
DDT
DDT,
chlordecone
DDT, PCBs,
chlordane,
lindane,
toxaphene,
heptachlor
Salmon
contaminated
with DDT,
PCBs, DDE,
mercury,
dioxin
2,3,7,8-
TCDD
SPECIES
Cells
Rat
Rat cells
Porcine
cells
Rat cells
Mouse
cells
Rat
Rat
EFFECTS
Disordered brain cell membranes
Decreases glycine levels in pons and
medulla
Binds to sodium channels, causing
persistent activation
Inhibits assembly of brain cell
tubulin
Inhibits ATPases involved in. ion
transport at nerve synapse
Stimulate protein kinase C
Increase behavioral reactions to
negative feeding events
Improper hypothalmic imprinting in
males
REFERENCE
Antunes-Madeira
& Madeira 1990
Truong el aL 1988
Lombet el aL
1988
Albertini el aL
1988
Kodavanti el aL
1988
Moser & Smart
1989
Daly 1991
Peterson 1992
70
-------
2.2.5.6 Endocrine Disruption
The endocrine system regulates physiological processes through a group of chemicals
called hormones, which are released by the endocrine organs and are transported via the blood
to other sites in the body where they exert their effect. They regulate responses to stress,
coordinate regulation of metabolism among muscle, liver, and fat, and coordinate function over
time, such as the changes required for normal sexual development and reproductive ability
(Hedge et aL 1987). Laboratory and field studies with freshwater and marine animals provide
evidence that xenobiotics are possibly contributing to the endocrine problems seen in the Great
Lakes, and other aquatic and marine systems. Effects from endocrine disruption such as thyroid
disorders, hormone deficiencies, secondary sex characteristic abnormalities, parental behavior
change, and hermapnroditism are found in many aquatic populations where elevated
concentrations of the chemicals of concern are found.
Wildlife Studies
No adult Great Lakes salmon (pink, coho, and Chinook) have been found without an
enlarged thyroid ("goiter") since 1974 by a team of researchers from Guelph University
(Leatherland 1992). Iodine deficiency was ruled out as a causal agent because Great Lakes fish
held comparable amounts of iodine to Northwest Pacific control fish. Thyroid enlargement and
reduced plasma thyroxin (T4) levels were induced in a dose-response manner in rats fed diets
of Great Lakes salmon, but were not inducible In fish fed the same diet (Leatherland 1992). No
contaminant analyses accompanied these findings.
Thyroid enlargement was also observed in the Great Lakes herring gulls in significantly
greater frequencies than in herring gulls from the Bay of Fundy (Moccia el aL 1986). Significant
differences were reported among and within lakes for the occurrence of increased thyroid mass
and thyroid tissue abnormalities, including epithelial cell hyperplasia, smaller follicular diameter,
taller epithelial cells, and less cellular colloid. Again, iodine deficiency was ruled out as a
causative agent. Exposure to environmental contaminants as a causative agent was supported by
geographic distribution of the effects as well as laboratory studies associating PCBs, DDT,
dieldrin, mirex, and heavy metals with the same thryoid anomalies (Moccia si aL 1986;
Government of Canada 1991). Fox and Peakall (1991) provided further evidence by
demonstrating an association between thyroid disorders and an environmental pollution gradient.
They also found that severity of goiter in Lake Ontario decreased hi subsequent collections, as
the contaminant load decreased, liver PCS level was significantly correlated with degree of
enlargement, and severity of thyroid enlargement was associated with retinoid depletion.
Other signals of endocrine disruption in salmon include premature sexual maturation while
never reaching full maturity (with loss of reproductive function accompanied by reduction in
expression of male hooked jaw and colored flanks), loss of sexual dimorphism (hermaphroditism
in males and females), low plasma estradiol and dihydroxyprogesterone levels, and low fertility
and embryo mortality resulting from low plasma steroid hormone levels (Moccia si aL 1981;
Leatherland si aL 1991; Leatherland 1992). Leatherland did not rule out genetic differences due
71
-------
to stock origin but suggested environmental agents as probable contributors to sexual precocity
and the loss of sexual dimorphism. For example, since 1980, the percentage of precocious coho
males in returning adults ranged from 40-60 percent in Lake Erie, whereas the percentage in
British Columbia (from the same genetic stock) ranged from 2-5 percent (Leatherland fit aL
1991). Lake Erie self-reproducing stocks also experience hermaphroditism. In other great
waters, between 29 percent and 55 percent of the burbot (Lota lota) collected on the north coast
of Bothnian Bay, Finland and Sweden, from 1987 to 1990 did not reach sexual maturity; between
87 percent to 98 percent near Tornio and Kemi were sterile (Pulliainen fit aL 1992). This sterility
was associated with irregular otolith growth and bone resorption. PCBs, DDT, dioxins, furans,
and metals were quantified. The decline hi striped bass from the San Francisco Bay delta was
attributed to reduced waterflow and increased xenobiotics affecting egg production and egg and
larval viability (Setzler-Hamilton et aL 1988). Reduced synthesis and resultant plasma/tissue
levels of sex hormones (estradiol, progesterone, and testosterone) have been associated with
elevated levels of cadmium, lead, BaP, PCBs, and mirex in sea stars (Asterias rubens), English
sole, Atlantic cod (Gadus morhua), Atlantic croaker, rainbow trout, polychaetes (Nereis virens),
and mussel (Voogt el aL 1987; Johnson el aL 1988; Freeman fit aL 1982; Thomas 1988; Chen
fit aL 1986; Fries and Lee 1984; Kluytmans el aL 1988). Dall's porpoises (Phocoenoides dalli)
from the northwest Pacific had reduced testosterone levels which were correlated with p,p'-DDE
concentrations (Subramanian el aL 1987). PCB and DDE exposure through diet caused a
reversible reduction hi retinol and thyroxin and failure of embryo implantation in harbor seals
(Brouwer fit aL 1989). Freeman and Sangalang (1977) studied the adrenal and testicular effects
of cadmium, arsenic, selenium, and Arochlor 1254 on grey seals (Halichoerus grypus). In this
study, all of these xenobiotics altered normal steroid biosynthesis.
Harbor seals from declining and stable populations of the Wadden Sea exhibited
significant reductions of plasma retinol and thyroid hormones (total and free thyroxin (T4), and
triiodothyronine (T3)) when fed a diet of PCB-contaminated Wadden Sea fish. A six-month diet
of relatively clean Atlantic mackerel (low PCBs) reversed the reduction. These field studies and
parallel laboratory studies led the researchers to suggest that reduced plasma retinol and thyroid
hormones from PCB exposure could increase susceptibility to infection by compromising the
seals' immune systems (Brouwer el aL 1989). PCBs hi the feral seals' fish diet were equivalent
to 25 (xg/kg body weight per day. The high-dose diet fed to confined seals was 1.5 mg PCB per
day and 0.4 mg p,p'-DDE and the low-dose was 0.22 mg PCB and 0.13 mg p,p'-DDE.
(Reijnders 1986; Brouwer fit aL 1989).
Little evidence of ovarian activity was reported by Beland el aL (1992) in female beluga
whales necropsied over the past 10 years. Thirty percent of the females were afollicular. Half
of the 19 to 25 year old females had mammary lesions. One out of 20 male specimens was a
true hermaphrodite.
Skewed sex ratios, reduced numBer of breeding males, female-pairing, and infertile
supernormal clutches have been observed hi Western and ring-billed gulls off the California
coast and Puget Sound, herring gulls of the Great Lakes, and U.S. Caspian terns (Hydroprogne
caspia) (Fox 1992; Fry el aL 1987; Fry and Toone 1981; Shugart el aL 1988). DDT and
72
-------
methoxychlor injected into gull (Larus californicus) eggs caused reproductive tract modification
of both sexes, and ovarian and oviduct tissue development in male embryos, effectively
feminizing the embryo (Fry el aL 1987; Fry and Toone 1981). Fox (1992) projected that the
feminization of male embryos from estrogenic agents such as DDT, mirex, TCDD, and
methoxychlor occurred during peak contamination years (1972-1976) in Lake Ontario and Lake
Michigan. Great Lakes herring gull endocrine disorders and reduced reproductive success
(embryo and chick mortality, edema, development abnormalities, and aberrant nesting behavior
such as female-female pairing) lessened with reduced contaminant levels (Gilbertson el aL 1991;
Fox 1992; Mineau el aL 1984; Peakall and Fox 1987). Caspian terns on the Great Lakes
continued to exhibit reduced reproductive success through the 80s, maintaining population levels
only through recruitment from less contaminated Canadian colonies (Fox 1992; Gilbertson el aL
1991).
Laboratory And Mechanistic Studies
The hormones of the endocrine system convey chemical signals to distant parts of the
body. Hormones influence cells by binding to specific cellular "receptors." Once bound to its
receptor, the hormone-receptor complex becomes activated, and will alter the cell's activity
(Figure 8). This is accomplished by influencing enzyme dynamics or inducing the expression
of specific genes. Gene products may be enzymes that modify the cell's metabolism, structural
proteins that will become part of the cell, or secretory materials. Hormones and their receptors
are therefore potent moderators of cellular structure and function.
Xenobiotics influence the endocrine system through several mechanisms. Hormone levels
in the blood can be affected by disruption or enhancement of their syntheses, and by increased
metabolic breakdown via the MFO system. Alternately, the cellular receptors of hormones may
be disrupted, making cells more or less responsive to hormonal signals. Dioxins are notorious
for influencing levels of endogenous receptors. TCDD modulates receptors for glucocorticoids,
prolactin, thyroxine, epidermal growth factor and estrogens (Umbreit and Gallo 1988).
This section will address xenobiotic effects on the endocrine system, including the thyroid,
adrenal gland and pancreas. The disruptive influence of xenobiotics on these glands and their
hormones is suspected to play a role in the wasting syndrome (Table 20). Xenobiotic effects on
reproductive hormones will be discussed later.
Effects on the Thyroid
The thyroid produces two hormones, thyroxine (T4) and triiodothyronine (T3), which are
involved in regulating cellular metabolism. Some of the xenobiotic substances known to affect
thyroid hormone levels are DDT, dioxin, PCBs, toxaphene and lead (Chu el aL 1986;
Tuppurainen el aL 1988). Disruption of thyroid homeostasis may be partly responsible for the
wasting syndrome. Xenobiotics can both decrease (hypothyroidism) and increase
(hyperthyroidism) thyroid activity, and, therefore, body metabolism. The effect observed depends
on the dose and duration of exposure. For example, DDT can both inhibit and stimulate thyroid
73
-------
activity, depending on dose. In pigeons, low doses of DDT produce hyperthyroidism, whereas
high doses cause hypothyroidism (Jefferies 1975).
TCDD has alternate effects on the two thyroid hormones. Although thyroxine levels in
the blood are depressed by TCDD, T3 levels are generally increased, although reports vary (Muzi
el aL 1987; Roth el aL 1988; Gorski el aL 1988b; Ivans el aL 1992). Thyroid stimulating
hormone (TSH) from the pituitary stimulates release of both T3 and T4. Slight alterations in TSH
levels have been reported following TCDD exposure (Henry and Gasiewicz 1987; Gorski el aL
1888a; Pohjanvirta el aL 1989a). However, the mechanism by which TCDD disrupts thyroid
hormone concentrations is still poorly understood (Roth el aL 1988).
TCDD-induccd alterations to thyroid hormones not only directly affect cell metabolism,
but can influence the overall body metabolism as well. Brown adipose tissue (which regulates
body temperature and weight through lipid and glucose metabolism) is secondarily affected by
TCDD-induced decreases in T4 (Weber el aL 1987; Rozman el aL 1987; Gorski el aL 1988b).
Unlike DDT and dioxin, PCBs and PBBs cause depression of both T3 and T4 levels in a
dose-related manner in mammals. Marmoset monkeys orally dosed with 0.1, 1.0, and 3.0
mg/kg/day PCB exhibited reduced serum T4 by 35, 81, and 99 percent, respectively (van den
Berg el aL 1988a). However, the effects in birds appeared to be related to the length of
exposure. PCB treatment of laying quail for 65-70 days resulted in depressed T4 and T3
concentrations, whereas prolonged exposure (120 days) increased plasma T4 values (Grassle and
Biessmann 1982).
The mechanism of PCB reduction in circulating thyroid hormones is two-fold. First, PCB
congeners reduce levels of thyroid hormones in the blood by having a strong affinity for T4
binding sites in prealbumin, the plasma transport protein for T4 (Rickenbacher el aL 1986).
Second, production of T3 and T4 in mammals is reduced due to direct damage to the thyroid
gland (Byrne el aL 1987; van den Berg el aL 1988a, b). There is not an increase in thyroid
hormone catabolism by the liver or other tissues (Byrne el aL 1987).
Other xenobiotic substances can also disrupt adrenal gland function. Toxaphene inhibited
corticosterone synthesis in the rat adrenal cortex (Mohammed el aL 1985). Veltman and Maines
(1986) found that 30 umol/kg mercuric chloride caused a 50 percent increase in MFO activity
in rat adrenal glands, causing subsequent disruption hi serum levels of adrenocortical hormones.
Effects on the Pancreas
Two hormones from the pancreas, insulin and glucagon, regulate glucose concentrations
in the blood. Hyperglycemia results from decreases in insulin, allowing blood sugar levels to
rise. The alternate, hypoglycemia, is due to decreased blood sugar. TCDD decreased insulin and
glucagon in rats (Gorski el aL 1988) and insulin in rabbits (Ebner el aL 1988), resulting in
transient hyperglycemia. In guinea pigs, insulin concentration was depressed for 10 days
following 1 mg/kg TCDD treatment (Brewster and Matsumura 1988). However, TCDD-induced
74
-------
FIGURE 8
MECHANISM OF HORMONE-RECEPTOR ACTION
Steroid
Hormone
Peptide
Hormone
Cytosol
Unactlvated Activated
Enzymes Enzymes
Biological
Responces
Binding to
Receptor
Activated
Receptor
Enzymes
Translation
Transcription
and other
Proteins
mRNA
f Biological
Responces
Mechanism of hormone action through cellular receptors. Peptide hormones attach to membrane-
bound receptors. The hormone-receptor complex activates enzymes, altering cellular processes.
Unlike peptide hormones, steroid hormones readily enter the cell. Once bound, the hormone-
receptor complex is activated and may interact with specific genes, inducing transcription to form
mRNA. The mRNA is translated in the cytosol to produce enzymes and other proteins, eliciting
a biological response.
-------
TABLE 20
MECHANISMS IMPLICATED IN THE WASTING SYNDROME
TARGET ORGAN
Liver
Brain
Thyroid
Adrenal Gland
Pancreas
MECHANISM
Mixed-function Oxidase
System
Neurotransmitters
Thyroxine & Triiodothyronine
Corticosterone
Insulin & Glucagon
PROCESS AFFECTED
Carbohydrate
Metabolism
Feeding Behavior
Cellular Metabolism
Brown Fat Metabolism
Gluconeogensis
Blood Glucose Levels
Effects on the Adrenal Glands
Corticosterone from the adrenal cortex is an important hormone in gluconeogenesis
(formation of new glucose molecules). Corticosterone levels were elevated 5-7 times normal
values in rats following TCDD treatment (Gorski el aL 1988a; Pohjanvirta el aL 1989a).
Adrenalectomy of rats drastically increased TCDD-induced mortality in rats (Gorski el aL
1988c), whereas corticosterone-replacement reduces mortality to nonadrenalectomized levels.
Corticosterone, therefore, provides partial protection from TCDD-induced toxicity in rats
resulting from reduced gluconeogenesis (Gorski el aL 1990).
Some of the effects of dioxins on adrenal hormones are mediated through receptor
disruption. TCDD treatment produces an approximately 30 percent decrease in binding capacities
of hepatic glucocorticoid receptors in female mice (Stohs el aL 1990; Lin el aL 1991b). This
effect does not appear to be regulated by the Ah locus. In rat liver, the dioxin and glucocorticoid
receptors are virtually indistinguishable physico-chemically (Cuthill el aL 1988).
Production of Corticosterone is controlled by adrenocorticotropic hormone (ACTH) from
the pituitary gland. Hypothysectomized rats suffer greater TCDD-induced toxicity, which is
returned to "normal" following administration of Corticosterone (Gorski el aL 1989d), suggesting
a role of ACTH in dioxin toxicity. However, alterations of serum Corticosterone levels are due
to altered responsiveness of the adrenal to ACTH simulation rather than to changes in plasma
ACTH levels (Jefcoate el aL 1987; DiBartolomeis el aL 1987; Moore el aL 1989). Kerkvliet el
aL (1990a) demonstrated that elevation of Corticosterone in mice exposed to either TCDD or
PCBs is dependent on the Ah receptor.
76
-------
hypoglycemia preceded insulin depression, indicating a period of insulin hypersensitivity (Gorski
and Rozman 1987). TCDD administration to rats further resulted in hypersensitivity to the
satiating effects of glucose and fructose (Pohjanvirta and Tuomisto 1990a). These effects on
pancreatic hormones may also play a role in the wasting syndrome by altering serum glucose
levels and peripheral satiety signals.
2.2.5.7 Reproductive Impairment
Wildilife Studies
A number of top predator species have exhibited reproductive problems or population
declines in a number of areas hi the Great Lakes basin since the 1950s. This list includes birds.
(the bald eagle (Haliaetus leucocephalus) (Postupalsky 1971a, b; IJC 1988), black-crowned
night-heron (Gilbertson personal communication 1988), Caspian tern (Kurita el aL 1987),
common tern (Gilbertson 1974a; Connors el aL 1975; Custer el aL 1988), double-crested
cormorant (Postupalsky 1976; Weseloh el aL 1983; Ludwig 1984), Forster's tern (Kubiak el aL
1989; Kubiak and Harris 1985), herring gull (Keith 1966; Ludwig and Tomoff 1966; Gilbertson
1974b; Mineau el aL 1984; Mineau and Weseloh 1981), osprey (Pandion haliaetus) (Berger and
Mueller n.d.; Postupalsky 1971a, 1980, 1983, 1985), and ring-billed gull (Sileo el aL 1977)),
mammals (the Beluga whale (Reeves and Mitchell 1984; Sergeant 1986; Beland el aL 1988;
Pippard 1985), mink, and otter (Lutra canadensis) (Pils 1987)); fish (the lake trout (Mac el aL
1985,1988)); and reptiles (the snapping turtle ((Chelydra serpentinafi (Brooks 1987). All of the
above animals depend upon Great Lakes fish for their food source. Researchers found relatively
high concentrations of organochlorine compounds, pesticides, and industrial chemicals in the
tissues of animals and their eggs in the affected populations (Ludwig and Tomoff 1966; Oilman
el aL 1977; Gilbertson and Fox 1977; Oilman el aL 1978; Frank el aL 1979; Haseltine el aL
1981; Hallett el aL 1982). Disorders which affect the success of reproduction hi the animals
included reduced fertility, reduced hatchability, reduced viability of offspring, impaired hormone
activity, or changed adult sexual behavior (described in the previous section Oh endocrine
disruption).
Common effects which characterize the current reproductive situation in the Great Lakes
are as follows:
• high tissue concentrations of PCBs, DDE, dieldrin, and/or other organochlorine
chemicals
• embryo toxicity and/or wasting
• offspring or embryo deformities
• adult parental behavioral change
• shoreline populations sparser than inland populations.
Scientific certainty in linking the observed effects with specific toxic chemicals has been
difficult due to the various analytical methods employed; numerous endpoints of effect; species,
age, and sex differences; and potential interactions between chemicals. Analagous evidence, such
77
-------
as observation of similar symptoms across a wide variety of organisms and contamination-linked
geographic locations, is often used to link contaminants with effects (Tillitt el aL 1992). In a
recent study which evaluated PCB residues in double-crested cormorant eggs, Tillitt elaL (1992)
statistically linked the observed reproductive effects (egg mortality) with PCBs measured as
dioxin equivalents (TCDD-EQ) using the H4IIE rat hepatoma cell bioassay. This study
demonstrated the relative enrichment in PCB potency hi the Great Lakes environment which may
explain 1) the observed variable reproductive success and 2) the continued adverse effects in the
populations, even though total PCBs have declined in the environment.
Eggshell thinning effects and accompanying reproductive loss as a result of DDT and its
metabolites were well-publicized in the 1960s and 1970s. As ambient levels of DDT declined,
many of the Great Lakes populations recovered. However, populations utilizing certain
geographical locations continue to exhibit reproductive failure (Peakall and Fox 1987; Peakall
1988; Fox el aL 1991; Harris 1988). In particular, areas of Lake Michigan, Lake Ontario, Lake
Superior, and Lake Huron remain affected by the contaminants of concern; Green Bay (Lake
Michigan), Saginaw Bay (Lake Huron), and Hamilton Harbor (Lake Ontario) are the most
influenced (Government of Canada 1991). Reproductive problems continue hi seven species of
Great Lakes birds, including the herring gull, ring-billed gull, common tern, Caspian tern,
Forster*s tem, black-crowned night-heron, bald eagle, double-crested cormorant, great blue heron
(Ardea herodias), and the Virginia rail (Ralus virginianus) (Government of Canada 1991).
Since 1980, double crested cormorants and ring-billed gulls numbers increased (Blokpoel
and Tessier 1986; Blokpoel 1988), although bald eagles, common terns, mink, and otters failed
to recover. Recent studies which compared Great Lakes inland versus shoreline bald eagle
populations found significantly lower reproductive success hi shoreline nests (Bowerman el aL
1991; Kubiak and Best 1991). The shoreline nests contained addled eggs with lethal
concentrations of PCBs, DDE, and dieldrin; 1987-1988 nestlings contained six tunes the PCB
and DDE plasma levels as did nestlings from the inland nests. Bald eagle productivity was
negatively correlated with PCB, DDE, and dieldrin load with the 1986-1990 breeding rate (0.6
young/occupied nest) too low to maintain a stable population (Bowerman el aJL 1991). Poor
Great Lakes shoreline reproduction or sparseness of populations has also been observed in
Forster's, common, and Caspian terns, mink, and river otters (Oilman el at, 1991; Government
of Canada 1991; Gilbertson el aL 1991). Correlations found between the hatching success of the
common snapping turtle and contaminated wetlands location between 1986 and 1989 demonstrate
the persistence of effects and locarional proximity (Bishop el aL 1991).
In order to maintain a stable bald eagle population, eagle eggs cannot exceed 3.5 ppm
DDE (Weimeyer el aL 1984), and, at 15 ppm DDE, populations of bald eagles suffer 100 percent
loss of productivity. Addled eggs collected in the Great Lakes basin between 1986 and 1990
held 3.4 to 20.5 ppm DDE (Kubiak and Best 1991) (Table 21).
Domestic mink fed Saginaw Bay carp contaminated with PCBs responded in a dose-
response manner in reproductive capability (number of offspring, kit body weight, and organ
weight) and kit survivability (Heaton el aL 1991). Wren el aL (1987) reported a syngergistic
78
-------
TABLE 21
MEASURES OF PRODUCTIVITY AND ADDLED EGG RESIDUES:
MICHIGAN, OHIO, AND ALASKA, 1986 - 1990
Lake Basin/Region
Addled Egg Residues1
(ug/g Fresh Wet Weight)
Productivity2
PCBs p,p'-DDE Dieldrin Prod. I3 Prod. 24
Lake Huron
Lake Michigan
Lake Erie
Lake Superior
Inland Ohio
Inland Mich.-U.P.
Inland Mich.-L.P.
Interior Alaska
76.7
41.0
22.1
10.1
10.7
7.5
8.2
1.4
20.5
20.1
3.4
4.5
1.9
3.2
2.7
0.5
1.16
1.32
0.43
0.25
0.19
0.24
0.11
0.02
0.59
0.68
0.75
0.84
0.71
0.93
1.14
1.29
41.2
48.0
52.6
55.4
57.1
59.7
71.8
76.8
1 Residues from 46 eggs collected from 36 breeding areas.
2 Productivities based on outcomes of 886 occupied breeding areas.
3 Number of fledged young per occupied breeding area.
4 Percent success rate of occupied breeding areas.
effect of methylmercury and PCS on mink kit growth and survival which exceeded the reduced
growth rate observed in kits exposed to 1.0 ug/g PCB in mothers' breast milk. These
experiments were conducted with mercury and PCB concentrations similar to those found in
some regions of the Great Lakes.
The reproductive success of the declining white croaker (Genyonemus lineatus) was shown
to be affected in spawning studies from a contaminated California site (San Pedro Bay) compared
to a reference site (Dana Point) (Hose el aL 1989). Ability to spawn, reduced fecundity (by 32
percent), reduced fertility (by 14 percent) and early oocyte loss (greater than 30 percent) were
associated with ovarian DDT concentrations. No fish with greater than 3.8 ppm DDT spawned;
79
-------
36 percent of the San Pedro sample had greater than 4 ppm ovarian DDT. Contaminant levels
(total DDT plus PCBs) in the sea-surface microlayer were found toxic to pelagic fish eggs and
larvae in this same area (Cross el aL 1987).
Mercury also impacts reproductive potential in both sexes. High rates of fetal mortality
result from in utero exposure during organogenesis (Eccles and Annau 1987). Pheasants treated
with mercury exhibited reduced egg production, hatchability and egg weight, and even production
of shell-less eggs (Fimreite 1971). Treatment of female mice with a single dose of
methylmercury resulted in increased losses in pre- and early post-implantation fetuses
(Verschaeve and Leonard 1984). Oral dosing of squirrel monkeys with 50 or 90 ng/kg
methylmercury for three months increased frequency of reproductive failure, decreased birth
weight and impaired offspring behavior (Burbacker el aL 1984). Mercury is present in breast
milk and crosses the placenta (Eccles and Annau 1987; Peterle 1991; Yoshida el aL 1992; Urbach
el aL 1992). Spermatogenesis is impaired in mice injected with 1 mg/kg methylmercury (Lee
and Dixon 1975). In vitro treatment of monkey sperm decreases sperm motility (Mohamed el
aL 1986a, b).
Kahn and Weis (1987) found differential resistance in the mummichog (Fundulus
heterclitus) from a mercury-polluted creek compared to a clean creek, as exhibited by reduced
fertility success attributed to changes in sperm motility. Inorganic mercury caused a significant
decrease in the fertility of the fish and offspring from the polluted creek, whereas highly toxic
methyl mercury (MeHg) did not. The reverse was seen in the control fish from the clean creek.
Susceptibility to inorganic mercury was attributed to the physiological cost of developing
pollutant tolerance, i.e., the inability to withstand further stress (Kahn and Weis 1987; Rahel
1981). Using sperm cell motility in the American sea urchin (Arbacia punctulata) as an index
for cell toxicity, Nelson (1990) demonstrated a biphasic dose-response in sperm motility
following exposure to paraoxon and dieldrin; sperm motility was inhibited by lindane; and
stimulated by mirex.
Reproductive effects in the endocrine system of marine animals have been associated with
heavy metals, atrazine, and chlorinated hydrocarbons such as PCBs, DDT, lindane, and
carbofuran (Sukumar and Karpagaganapathy 1992; Reijnders and Brasseur 1992; Reijnders 1986;
Simic el aL 1991; Batty 1990). Carbofuran exposure resulted in atretic oocytes, retrogressive
ovaries, oocyte-depleted germinal vesicles, and reduced yolk granules in fresh-water fish (Colisa
alia) (Sukumar and Karpagaganapathy 1992). Uterine occlusions and stenoses, bilateral
adrenocortical hyperplasia, and hormonal osteoporosis observed in pinnipeds were associated with
PCBs and DDT (Baker 1989; Bergman and Olsson 1985; Brouwer el aL 1989; Helle el aL 1976a,
b; Reijnders 1986; Reijnders and Brasseur 1992). Cadmium, lead, and PCBs have affected
biosynthesis of reproductive hormones in other marine animals as described in the previous
section on endocrine disruption (den Besten 1991; Freeman el aL 1982; Johnson el aL 1988;
Voogt el aL 1987; Thomas 1988).
80
-------
Information regarding contaminant effects on humans are limited primarily to studies of
contamination from occupational disasters, cohort studies, and clinical reports described in
Section 2.2.5.8 (Fein el aL 1984; Rogan ei aL 1986; Rogan el aL 1988; Jacobson el aL 1990;
Jacobson and Jacobson 1991; Leoni el aL 1989; Bush ei aL 1986). In recent years, a number of
studies have linked reproductive changes in humans with ambient exposure. For example,
findings from studies of the Michigan Maternal/Infant Cohort associated reproductive effects (l°w
birth weight, shorter gestional age, smaller head circumference) with the lifetime experience of
the mothers' Lake Michigan fish consumption (Jacobson el aL 1990; Jacobson and Jacobson
1991). Bush and coworkers (1986) found an association between the presence of three PCB
congeners: (2,3,4,2'A',5'- IUPAC No. 153,2,3,5,2',3',4'- IUPAC No. 137 and 2,4,5,3',4'- IUPAC
No. 123 and loss of sperm motility in males with fertility problems. Carlsen el aL (1992), in a
meta-analysis of sperm count studies dating back to 1938, found an approximate 50 percent
reduction in sperm count and a significant decrease in seminal fluid volume in men worldwide
between 1938 and 1991. Genetic changes were ruled out since the change was worldwide over
one generation. Among his suggestions for why sperm numbers have declined, Sharpe (1992)
points out that exposure to DDT, PCB, and other chemicals capable of disrupting the endocrine
system during a critical window of time in early intra-uterine development can affect the
production of spermatogonia. This hypothesis is supported by the timing of the chemical
revolution since World War II and the concomitant decrease in sperm count in male humans.
Laboratory And Mechanistic Studies
Some of the most insidious effects of airborne water pollutants are those on reproduction.
Reproductive impairments are largely due to endocrine disruption. Xenobiotic compounds can
affect endocrine regulation of reproduction by a variety of means, including disrupting pituitary
control of reproductive cycles, altering metabolic synthesis or breakdown of hormones,
mimicking natural endogenous hormones, and antagonizing or blocking hormonal signals.
The levels of steroids in both males and females, as well as their reproductive cycles, are
regulated by peptide hormones, "such as luteinizing hormone (LH) from the hypothalamus. LH
stimulates production of sex steroids (estrogen, progesterone, and testosterone) by the gonads,
and is regulated by gonadotropin-releasing hormone (GnRH) from the hypothalamus. Males and
females exhibit differences in their pattern of LH secretion. Females release LH in a pulsatile
manner and exhibit a surge of LH secretion that stimulates ovulation of eggs from the ovary.
Males produce relatively constant quantities of LH, and are not capable of producing an LH
surge. These patterns are established during embryonic development, or shortly following birth.
Some toxic substances can drastically alter reproductive function by disrupting LH
secretion from the pituitary, thereby upsetting the reproductive regulatory center. TCDD
decreases GnRH receptors in the pituitary of male rats, thereby reducing the piruitary's
responsiveness to androgen deficiency and preventing compensatory increases in LH secretion
(Bookstaff el aL 1990). Other compounds, such as DDT, DDE and parathion, also decrease LH
levels in adults (Gellert el aL 1972; Richie and Peterle 1979; Rartner el aL 1984; Rattner el aL
1982a, .b; Rattner and Ottinger 1984). Single dose exposure of pregnant mice to 0.16 fim/kg
81
-------
TCDD feminize LH secretory patterns in her male offspring as adults (Mably el aL 1992). By
altering levels of LH, or its pattern of secretion, xenobiotics significantly impair reproduction in
both males and females.
Another pituitary hormone involved with reproduction is prolactin, which stimulates
production of milk in female mammals and influences reproductive functions in other vertebrate
groups. Many different xenobiotics have been demonstrated to disrupt serum prolactin
concentrations. Prolactin levels were altered in ring doves fed diets containing a mixture of DDE,
PCB, and mirex (McArthur el aL 1983). TCDD significantly reduces serum prolactin
concentrations in rats within 4 hours of treatment (Jones el aL 1987; Russell el aL 1988; Moore
el aL 1989). This effect is correlated with a dramatic increase hi dopamine hi the brain (Russell
el aL 1988). Orcadian alterations of prolactin secretion (Jones el aL 1987) may be influenced
by TCDD-induced alterations in melatonin release (Linden el aL 1991). TCDD also alters levels
of prolactin receptors hi many tissues. Seven days following TCDD treatment, hepatic prolactin
receptors are reduced by 78 percent hi liver, but increased to 191 percent in kidney (Jones el aL
1987).
Xenobiotic compounds can alter levels of endogenous hormones. PCBs disrupt levels of
the pregnancy-maintaining hormone progesterone hi monkeys (Truelove el aL 1990). PCBs also
cause increased levels of estrogens and prostaglandins during pregnancy (Lundkvist and Kindahl
1989). Androgen deficiency induced by TCDD treatment hi rats may be the result of a decrease
hi testosterone secretion by the testicles (Moore and Peterson 1988).
An important mechanism for altering steroid hormone levels is through the MFO system.
Several MFO enzymes are involved hi the biosynthesis of sex steroids (Table 22). All steroids
are derived from cholesterol, and many serve as substrates for the formation of others. For
instance, the female steroid progesterone is utilized by males to make testosterone, and females
use testosterone as a necessary building block for estrogens. Other MFOs eliminate sex steroids
by oxidizing them to forms readily excretable by the kidneys. The MFO system is, therefore,
integral in the regulation of sex-steroid levels hi the blood, either by their synthesis,
interconversion of one form to another, or by metabolism into waste products that are eliminated
from the body. By inducing the MFO system, xenobiotics are able to drastically alter levels of
sex steroids hi the body (Dieringer el aL 1979; Truscott el aL 1983; Gustafsson el aL 1983; Khan
1984; Payne el aL 1987). Xenobiotics may induce some MFO enzymes but inhibit others
(Voorman and Aust 1987, 1989). The inhibition of estradiol hydroxylase activity by TCDD
(Voorman and Aust 1989) may help explain the TCDD-induced increase hi estrogen levels
(Gallo 1988). Examples of MFO induction and its reproductive effects either by hydrocarbons
or specific xenobiotics are presented in Table 23 and Table 24, respectively.
Many xenobiotics mimic natural hormones. DDT is an artificial estrogen, and probably
the best studied example of an exogenous hormone mimic (Bulger and Kupfer 1983; McLachlan
1985). The earliest laboratory account of the estrogenic nature of DDT was the discovery that
DDT was uterotropic (increased uterine weight) in rats (Leven el aL 1968; Welch el aL 1969).
Further, mice exposed to DDT exhibited prolonged estrous cycles and decreases in ova
82
-------
implantation (Lundberg 1973). It was subsequently established that the o,p'-isomer of DDT was
largely responsible for the uterotropic activity (Welch el aL 1969). DDT binds to the cellular
estrogen receptor and initiates the same sequence of events as natural estrogen (Nelson 1974),
including an increase uterine DNA synthesis (Ireland el aL 1980) and induction of protein
synthesis and secretion (Stancel el aL 1980). Many of these induced proteins are enzymatic in
nature (Singhal el aL 1970; Cohen el aL 1970; Kaye el aL 1971; Bulger el aL 1978b; Bulger and
Kupfer 1978, 1983b). Particularly notable, one of the proteins induced by o,p'-DDT in the rat
uterus is the receptor molecule for another sex steroid, progesterone (Mason and Schulte 1980).
Other xenobiotics are also hormone mimics. PCBs have extensive effects on reproductive
systems (Reijnders 1988), including stimulation of uterine weight increases, prolonged estrous
cycles, unpaired fertility, reduced number of young, and reduced maternal ability to carry young
to term (Table 25). These effects are mediated in part by PCBs ability to bind to uterine estrogen
receptors (Korach el aL 1988). PCBs also bind to other receptors hi the rat liver (Buff and
Brundl 1992), possibly interfering with the function of these endogenous receptors, which also
bind the thyroid hormones thyroxine and triiodothyronine. Some of TCDD's estrogenic properties
may be due to its ability to bind to estrogen receptors (Umbreit el aL 1989b).
Some xenobiotics only mimic endogenous hormones after being metabolized, or activated,
in the body. Methoxychlor (bis-p-methoxy DDT) is a proestrogen and is metabolized by the
hepatic MFO system into estrogenic products (Nelson el aL 1976, 1978; Budger el aL 1978c;
Ousterhout el aL 1979, 1981). The estrogenic metabolite of methoxychlor (HPTE) was shown
to be about 10 times more active than o,p'-DDT (Ousterhout el aL 1981). See Table 26 for the
estrogenic effects of methoxychlor on reproduction.
Xenobiotics may also block or reduce the activity of endogenous hormones. Many of
these have antiestrogenic effects in females, such as a decrease in: 1) uterine weight, 2) cell
growth, 3) estrogen-induced protein secretion, 4) estrogen and progesterone receptors, 5)
peroxidase activity, 6) estrogen-stimulated c-fos oncogene mRNA, 7) epidermal growth factor
receptor binding activity, and 8) EOF mRNA levels (Table 27). Antiestrogenic compounds can
impair female reproductive capacity, including the ability to conceive, maintain young throughout
pregnancy, deliver, and care for young postnatally.
There are several mechanisms for these antiestrogenic effects. TCDD directly reduces the
concentration of estradiol-176 in human tissues by increasing the metabolism of estradiol to a
less active form (Graham el aL 1988; Gierthy el aL 1987; Spink el aL 1990; Spink el aL 1992).
The antiestrogenic effect of TCDD in many cases is mediated not by reductions in estrogen, but
by its ability to down-regulate estrogen receptors (Romkes el aL 1987; Umbreit and Gallo 1988;
DeVito el aL 1992). Ten nM TCDD can cause up to a 74 percent decrease hi estrogen receptor
levels hi mouse cells in 6 .hours (Zacharewski el aL 1991), and a 63 percent decrease in human
cells by 12 hours following treatment (Harris el aL 1990). Estrogen receptor down-regulation
is dependent upon dioxin binding to the Ah receptor (Gasiewicz and Rucci 1991). In normal
circumstances, estradiol mediates some of its effects through small, regulatory proteins called
growth factors. For example, estradiol induces receptors for epidermal growth factor (EGF).
83
-------
TCDD inhibits estradiol's induction of EOF receptors (Astroff el aL 1990; Safe el aL 1991; Abbot
el aL 1992). The TCDD-induced decreases in both estrogen and growth factor receptors are
mediated through the aryl hydrocarbon (Ah) receptor (Zacharewski el aL 1991, 1992; Lin si aL
1991a, b; Abbot el aL 1992; Schrenck el aL 1992).
In males, xenobiotics may exhibit estrogenic or antiandrogenic activities (Table 28).
Effects include testicular atrophy, reduced fertility and arrested spermatogenesis. Reduced levels
of androgens are related to both decreased secretion from the testes and increased metabolism
via induction of the MFO system.
TABLE 22
STEROID HORMONE SYNTHESIS BY MIXED-FUNCTION OXTOASES
CYTOCHROME
P450W
P45017llp)tt
P450JQX family
(Aromatase system)
P45021 and
P450llbett
SUBSTRATE
Cholesterol
Progesterone
Testosterone
Progesterone
PRODUCT
Pregnenolone
Testosterone
Estradiol
Cortisol, Corticosterone,
and Aldosterone
Source: Fevold 1983; Nebert and Gonzalez 1987; Simpson and Waterman 1989
84
-------
TABLE 23
EFFECTS OF HYDROCARBONS ON MFO INDUCTION
AND REPRODUCTIVE IMPAIRMENT
SPECIES
Gunners
Chicken
Herring
gulls
Salmon
Flounder
Herring
gulls
Mallard
ducks
Trout
EFFECTS
No evidence for altered steroid
metabolism
MFO induction in kidney
MFO induction in kidney
Increased levels of sex steroids in bile
Inverse relationship between MFO
induction and fertilization success
MFO induction
MFO induction
No evidence for reproductive
impairment
REFERENCE
Hellou & Payne 1986
Lee el aL 1986
Lee el aL 1985
Truscott el aL 1984
Spies el aL 1984
Gorsline si aL 1981
Miller el aL 1978
Hodgins el aL 1977
85
-------
TABLE 24
EFFECTS SPECIFIC XENOBIOTICS
ON MFC INDUCTION AND REPRODUCTIVE IMPAIRMENT
COMPOUND
TCDD
TCDD
TCDD
DDT
PCBs
PCBs
PBBs
HCB
Mercury
SPECIES
Rat
Rat
Rat
Rat
Pigeon
Salmon &
Flounder
Rat
Rat
Rat
EFFECTS
Decreased plasma testosterone
and dihydrotestosterone by 90
percent and 75 percent,
respectively
Decreased estradiol
Decreased androgen
concentrations, reduced sex
glands and reproductive capacity
Induced MFO enzymes that
metabolize androgens
Induced several P450 isofonns
Decreased androgen
concentration
Increased steroid catabolism
Induced MFO enzymes that
metabolize androgens
Induced of MFOs and alteration
of adrenal steroid metabolism
REFERENCE
Moore el aL 1985
Gierthy el aL
1987
Sager 1983
Haake el aL 1987
Borlakoglu el aL
1991
Truescott el aL
1983
McCormack el aL
1979
Haake el aL 1987
Veltman and
Maines 1986
86
-------
TABLE 25
REPRODUCTIVE EFFECTS OF POLYCHLORINATED BIPHENYLS
SPECIES
Rhesus
monkey
Guinea
Pig
Marmoset
monkey
Mourning
dove
Mink
Japanese
quail
Mink
Rhesus
Monkey
Rat
Mouse
Fish
EFFECTS
Altered progesterone levels and
increased duration of menses
Increased levels of estrogens and
prostaglandins
Absence of corpora lutea
Altered progesterone levels and reduced
reproductive success
Decreased number of young
Decreased plasma estradiol levels before
sexual maturity, delayed oviposition and
diminished laying capacity
Decreased number of young
Impaired fertility and ability to carry
infants to term
Decreased number of young
Prolonged estrous cycle
Reabsorption of egg sac
REFERENCE
Truelove el aL 1990
Lundkvist and Kindahl
1989
van den Berg el aL
1988b
Koval el aL 1987
den Boer 1983
Biessmann 1982
Jensen el aL 1977
Allen and Barsotti
1976
Under el aL 1974
Orberg and Kihlstroem
1973
Mac el aL 1988
87
-------
TABLE 26
REPRODUCTIVE EFFECTS OF METHOXYCHLOR
SPECIES
Mouse
Mouse
Mouse
Mouse
Rat and
Hamster
Cells
Rat
Rat
Rat
Mouse
Rat
EFFECTS
Induced steroid secretion by ovarian cells
Stimulation of uterus & its secretions
indistinguishable from that of estradiol
Stimulated uterine hypertrophy
Increased uterine weight
Induced behavioral estrus
Metabolites of methoxychlor are potent
estrogens
Methoxychior is a proestrogen
Methoxychlor binds to uterine estrogen
receptors
Methoxychlor 16 times less estrogenic
than o,p'- DDT
Methoxychlor is a proestrogen
Increased uterine weight
REFERENCE
Martinez and Swartz
1992
Rourke el aL 1991
Eroschenko 1991
Eroschenko and
Cooke 1990
Gray el aL 1988
Kupfer and Bulger
1987
Bulger el aL 1978c
Nelson 1974
Bitman and Cecil
1970
Kapoor el aL 1970
Welch el aL 1969
-------
TABLE 27
ANTEESTROGENIC EFFECTS OF XENOBIOTICS IN FEMALES
COMPOUND
TCDD
TCDD
TCDD
TCDD
TCDD
TCDD
TCDD
TCDD
TCDD
TCDD
Lindane
Lindane
SPECIES
Rat
Mouse
Rat
Rat
Rat
Human
cells
Rat
Hamster,
guinea pig,
rat
Mouse
Pike
Rat
Rat
EFFECTS
Decreased: uterine weight; estrogen &
progesterone receptors; EOF binding;
and enzyme activity
Inhibited estrogen-induced EGF
receptors
Decreased: uterine weight; estrogen &
progesterone receptors; EGF binding
and receptors; c-fos mRNA levels;
. and enzyme activity
Decreased: uterine weight; estrogen,
progesterone and EGF receptors; EGF
mRNA levels; and enzyme activity
Decreased c-fos mRNA levels
Altered secretion of estrogen-induced
proteins
Decreased uterine EGF receptor
binding activity and EGF receptor
mRNA
Altered estrogen metabolism
Depressed estrogen-induced uterine
weight gain
Retarded egg development and fry
growth
Delayed vaginal opening, disrupted
cycles, reduced uterine weight
Ovarian atrophy and unpaired
oogenesis
REFERENCE
Dickerson si aL
1992
Abbot el al
1992
Safe el aL 1991
Astroff and Safe
1991
Astroff ei aL
1991
Biegel & Safe
1990
Astroff el aL
1990
Umbreit el aL
1989a
Umbreit el aL
1988
Helder 1980
Chadwick el aL
1988
van Giersbergeh
el aL 1986
-------
In some cases, the mechanism of action remains obscure, even after extensive research.
An example is the effect of DDE (an analog of the pesticide DDT) on eggshell thickness in birds.
Ratcliffe (1967) was the first to report the toxic effects of substances on eggshell weights.
Mallard hens fed 50 ppm DDT produced eggshells that were 18 percent thinner and weighed 12
percent less (Kolaja and Hinten 1979). Both alteration in metabolism of steroids (Peakall 1967;
1970a, b; Lustick el aL 1973; Peterle el aL 1974; Haegele and Tucker 1974) and impairment of
steroid binding to cellular receptors (Lundholm 1987) have been reported hi birds exposed to
DDE. Alterations in levels of parathyroid hormone (which is involved in regulating calcium
concentrations) may be involved in eggshell thinning (Parsons and Peterle 1977; Haseltine el aL
1981). DDT and DDE also are potent inhibitors of calmodulin, a cellular protein important for
proper deposition of eggshell calcium (Lundholm 1987). However, in spite of intensive
investigation, the exact mechanism by which DDE reduces eggshell thickness is still poorly
understood (Peterle 1991).
Xenobiotic contaminants cause numerous effects on developing young (see
Transgenerational Effects, Section 2.2.5.8). Xenobiotics both cross the placental barrier (van den
Berg el al. 1987) and are transferred to newboms via breast milk (Courtney and Andrews 1985).
In pheasants, 1 percent of TCDD administered to the female is incorporated into each of her first
15 eggs (Nosek el aL 1992). Further, TCDD is known to reduce transfer of placental nutrients
to developing young (Manchester el aL 1987), thereby impairing development.
2.2.5.8 Transgenerational Effects
An increasing body of evidence describing the effects of low-level, chronic exposure to
twentieth century chemicals has caused lexicologists to expand their perspective of concern from
impacts on the exposed organism to consideration of effects on the progeny born to the originally
exposed individual. In many cases, the parent organism is apparently unaffected by the exposure,
but serves only as an accumulator of contaminants, ultimately exposing the offspring where an
effect may occur. The health impacts resulting from the exposure of progeny secondarily to the
original parentally acquired contaminants are referred to as a transgenerational effects. In
humans, this secondary exposure of the progeny can take two forms: (1) in utero exposure prior
to parturition or hatching, and (2) postpartum exposure of the newborn via breast milk.
Approximately 25 chemical substances are known to produce transgenerational effects in
humans, while over 800 are known to do so in laboratory animals (Kurzel and Cetrulo 1981).
The reasons for this discrepancy include both the fact that humans are more resistant to some of
these substances, and that subtle alterations or deficits in neuromuscular maturity, body weight,
physical size, autonomic regulation, behavioral endpoints, and the like have only recently begun
to be investigated (Fein el aL 1983; Jacobson el aL 1992).
With respect to in utero exposure of the human, there are three developmental periods
during which the unborn child is at risk of impairment (Kurzel and Cetrulo 1981). These
developmental periods, summarized in Table 29, are: (1) fertilization and-implantation, (2) the
embryonic period, and (3). the period of fetal development.
90
-------
I
TABLE 28
iSTROGENIC AND ANTIANDROGENIC EFFECTS OF XENOBIOTICS IN
MALES
COMPOUND
TCDD
TCDD
PCB
DDT
DDT and
Methoxychlor
DDT
Lindane
Lindane
Lindane
HCB
SPECIES
Rat
Rat
Rat
Rat
Rat
Rat
Rat
Rat
Rat
Rat
EFFECTS
Decreased androgen
secretion
Reduced testosterone 90
percent, dihydrotestosterone
75 percent, and reduced
testis and epididymis
weights
Increased testis weight
Induce MFO enzymes that
metabolize androgens
Bind to testicular estrogen
receptors
Blocks androgen binding to
prostate receptors
Inhibited spermatogenesis,
seminiferous tubules
degenerated
Estrogenic effect, including
atrophic testes and
spennatogenic arrest
Estrogenic effect
Induce MFO enzymes that
metabolize androgens
REFERENCE
Moore and
Peterson 1988
Moore el al.
1985
Johansson
1987
Haake elaL
1987
Bulger el aL
1978a
Wakeling and
Visek 1973
Chowdhury el
311987
van Velson
el aL 1986
van Giersbergen
el aL 1984
Haake el aL
1987
-------
Aside from the small percentage of morphologic abnormalities, or birth defects,
attributable to chemical contaminants — estimated to be 4-6 percent of all birth defects (Kurzel
and Cetrulo 1981) — the majority of the observed effects will be associated with the fetal
development period.
In this period, toxic effects are usually manifested in a diminution of cell size or a
reduction in cell numbers. Since this developmental phase represents a period of unprecedented
growth and maturation of tissues (Calabrese and Sorenson 1977), growth retardation and
functional deficits, including central nervous system injury or retarded development, usually result
from insult during this stage of development. The developing brain and central nervous system
are particularly susceptible to impact, since development processes, including myelination, are
not complete, even at birth. Further, the developing fetus is likely to be more susceptible to
insult by toxic substances because of the incomplete development of its liver enzyme systems,
and a relatively poorly developed blood-brain barrier (Calabrese and Sorenson 1977).
TABLE 29
EFFECTS OF CHEMICAL EXPOSURE DURING HUMAN DEVELOPMENTAL
PERIODS ASSOCIATED WITH INTRAUTERINE LIFE
Functional
Period
Fertilization
and
Implantation
Embryonic
Development
Fetal
Development
Intrauterine
Time Period
Conception -
17 days
18-55 days
56 days - Term
Developmental
Stage
Primary germ cells;
blastocyst; gastrula
Organogenesis
Growth; maturation
of tissues; several
differentiations
Developmental
Decrement
Cell death - alternative cells
recover and multiply;
organism death with abortion or
reabsorption
Morphologic or organ
system abnormalities
Growth retardation;
functional deficits
Source: Developed from the data of Kurzel and Cetnilo (1981).
92
-------
A variety of toxic compounds are capable of being transplacentally transmitted from
human mother to fetus, and an even larger array of substances can be transferred from mother
to newborn in breast milk. Among those substances transferred transplacentally are cadmium
(Korpela el aL 1986; Bonithon-Kopp el aL 1986; Lauwerys 1986), lead (Korpela el aL 1986;
Bonithon-Kopp el aL 1986; Li 1988), mercury (Bonithon-Kopp el aL 1986; Harada 1977;
Takeuchi 1972; Spencer el aL 1988), hexachlorobenzene (Bush el aL 1984), metabolites of DDT
(Rogan el aL 1986b), dieldrin (Colborn 1989), and polychlorinated biphenyl (PCB) (Rogan el aL
1988; Rogan el aL 1986b; Bush el aL 1984; Jacobson el aL 1983; Kodama and Ota 1980; Masuda
el aL 1978; Polishuk el aL 1977; Funatsu el aL 1972). Among the contaminants potentially
transferred from mother to infant in breast milk are cadmium (Dabeka el aL 1986; Stemowsky
and Wessolowski 1985), lead (Stemowsky and Wessolowski 1985), mercury (Colbom 1989),
hexachlorobenzene (Mes el aL 1984; Mes and Davies 1979), metabolites of DDT (Rogan el aL
1987; Davies and Mes 1987; Rogan el aL 1986a; Mes el aL 1986; Mes el aL 1984; Cone el aL
1983; Mes and Davies 1979), dieldrin (Davies and Mes 1987; Mes el aL 1986; Mes el aL 1984;
Mes and Davies 1979), hexachlorocyclohexane (Davies and Mes 1987; Mes el aL 1986; Mes el
aL 1984; Mes and Davies 1979), heptachlor epoxide (Mes el aL 1986; Mes el aL 1984; Mes and
Davies 1979), chlordane fractions, including oxychlordane and trans-nonachlor (Davies and Mes
1987; Mes el aL 1984; Mes and Davies 1979), photomirex (Davies and Mes 1987; Mes el aL
1986), and polychlorinated biphenyls (Rogan el aL 1987; Mes el aL 1987; Rogan el aL 1986a,
b; Mes el aL 1986; Mes el aL 1984; Cone el aL 1983; Wickizer el aL 1981; Mes and Davies
1979; Grant el aL 1976).
The concept of transgenerational effects resulting from exposure to an exogenous chemical
compound is not new. Traditional teratology has frequently associated morphologic alterations
and physical malformations in the embryo or fetus with the impacts of in utero exposure to
external dismissal agents. Classic examples are to be found in association with known
administration of prescription drugs, e.g., limb deformities associated with maternal dosages of
thalidomide during pregnancy (Tuchmann-Duplessis 1975), and genital anomalies associated with
maternal ingestion of diethylstilbestrol (DBS) to prevent miscarriages (Kurzel and Cetrulo 1981).
Additional evidence is provided from a considerable body of knowledge developed from research
on the use of "recreational drugs", e.g., craniofacial anomalies associated with fetal alcohol
syndrome (Able 1984; Jones el aL 1973), and reduced head circumference and body size of
infants who were exposed to nicotine as a result of maternal smoking (USPHS 1979).
•
Only recently, however, have investigations been oriented toward the more subtle
transgenerational effects of exogenous chemical substances. Some of these studies have been
oriented toward chemical substances to which the mother was deliberately exposed, e.g., alcohol
(Coles el aL 1985; Golden el aL 1982; Streissguth el aL 1980, 1983, 1984), marijuana (Fried
1982), cocaine (Chasnoff el aL 1985), and methadone (Hans el aL 1984). Other studies
considered the effects of inadvertent maternal exposures, chiefly to environmental contaminants,
e.g., lead (Bellinger el aL 1987; Ernhart el aL 1987; Dietrich el aL 1986), mercury (Harada 1976;
Takeuchi 1972a, b), and polychlorinated biphenyls (Jacobson el aL 1985; Rogan el aL 1986a).
93
-------
\ I
From these studies of subtle effects resulting from transgenerational exposures to
exogenous chemical substances, i.e., effects other than physical dysmorphology, a series of
principles have emerged. These include:
1. Transgenerational effects are negative, frequently subtle, and diminish the potential
of the impacted offspring, either physically, behaviorally, emotionally, cognitively
or in some combination of these factors, (Rogan el aL 1988, 1986a, 1986b;
Jacobson el aL 1990a, b, 1985, 1984a).
2. Exposure to exogenous chemical substances which may produce asymptomatic,
sub-clinical, or no apparent effects in the pregnant mother, may have profound
effects upon the embryo or fetus (Takeuchi 1972b; Jacobson el aL 1985; Rogan
el aL 1986a; Rogan fit aL 1988).
3. If maternal effects are observed as a result of exposure, the sequelae observed in
infants bom to .these mothers may differ significantly both hi character and degree
(Takeuchi 1972a; Funatsu and Yamashita 1972).
4. The deficits produced in transgenerationally exposed offspring are usually durable,
i.e., of a long-lasting nature, frequently persisting a life-time (Jacobson el aL
1990a, b; Rogan el aL 1988; Harada 1977; Takeuchi 1972b).
5. Transgenerational exposure may result in clinically normal newboms whose long-
term deficits are not evident until later in life (Jacobson el aL 1990; Jacobson el
aL 1989).
6. Not only is the extent and duration of exposure important to the degree or
magnitude of the effect observed, but the tuning of the exposure is critical to the
character and potential of the adverse outcome (Kurzel and Cetrulo 1981;
Jacobson el aL 1989; Jacobson fit aL 1990; Harada 1976).
7. Profound transgenerational effects may result from either an acute, single maternal
exposure (Rogan el aL 1988; Rogan 1982; Wong and Huang 1981; Harada 1976;
Higuchi 1976), or, because of the excessive biological half-lives of some of these
compounds (Bush el aL 1984), transgenerational effects may result from small,
cumulative exposures over an extended period of tune (Jacobson el aL 1990a, b,
1984a, b; Rogan el aL 1986a, b).
8. Because of the excessive biological half-lives of some of these compounds and
then* storage hi maternal tissues, transgenerational effects in progeny may occur
in association with pregnancies occurring years after maternal exposure has ceased
(Harada 1976; Abe el aL 1975).
94
-------
9. Because of the extensive biological half-lives of some of these compounds, there
is a potential for multi-generational effects, i.e., a single maternal exposure may
effect more than one generation of the progeny born to that mother (Swain 1988).
Now, as never before, the developing body of knowledge related to transgenerational
effects has underscored the need to evaluate the safety of chemicals never intended for human
consumption.
22.6 Case Studies of Multiple Effects
of Compounds of Concern
•
2.2.6.1 Adverse Consequences of Eutrophication in
Estuaries And Coastal Seas
Although nitrogen and phosphorus are essential for plant growth, excesses of these
nutrients produce severely negative impacts on aquatic and marine ecosystems. Among these
deleterious effects are hypoxia, anoxia, reduction of plant biomass, and the proliferation of
nuisance algae blooms. These negative consequences of eutrophication are discussed in the
following case study.
Anoxia And Hypoxia
Anoxia is the complete removal of dissolved oxygen from the water column, an event
which obviously causes widespread damage to aquatic plants and animals. Even mobile animals
which can escape from anoxic waters can suffer population declines from the loss of habitat area.
For example, in parts of the Baltic Sea cod eggs laid in oxic surface waters sink into anoxic
bottom waters where they die (Rosenberg el aL 1990). Oxygen concentrations in the bottom
waters of the deep basins of the Baltic between 1969 and 1983 are correlated with codfish
populations (Hansson and Rudstam 1990). Price el aL (1985) have speculated that the decline
of striped bass populations in part of Chesapeake Bay may be a result of the increasing volume
of anoxic bottom waters; the striped bass have been forced into more shallow and warmer waters,
waters which may in fact be excessively warm for this species to thrive.
Oxygen need not be completely absent for damage to occur, and a lowering of oxygen
to concentrations as low as 3 to 4.3 mg liter"1 can cause ecological harm in some estuaries and
coastal seas (EPA 1991). Such a depletion of oxygen is termed hypoxia. Examples of ecological
damage from hypoxia include lowered survival of larval fish, mortality of some species of
benthic invertebrates, and loss of habitat for some mobile species of fish and shellfish which
require higher concentrations of oxygen, such as lobster and codfish (Baden el aL 1990; EPA
1991). Significant mortalities of lobsters and population declines of both lobster and codfish
have been observed in some Swedish coastal waters as a result of increased incidences of
hypoxia (Baden el aL 1990).
95
-------
Anoxia and hypoxia are major and growing problems in many estuaries and coastal seas.
Over the past few decades, the volume of anoxic bottom waters has been increasing in
Chesapeake Bay (Officer el aL 1984; D"Elia 1987), the Baltic Sea (Larsson el aL 1985), and the
Black Sea (Lein and Ivanov 1992). The apex of the New York Bight (an area of some 1,250
km2) becomes hypoxic every year, and a large region of the Bight became anoxic in 1976
(Mearos el aJL 1982). Hypoxic events appear to be becoming more common in waters such as
Long Island Sound (EPA 1991; Parker and O'Reilly 1991), the North Sea (Rosenberg 1985), and
the Kattegat (the waters between Denmark and Sweden; Baden el aL 1990), although historical
data on oxygen concentrations hi coastal waters are often poor.
Anoxia and hypoxia result from oxygen consumption exceeding oxygen supply. Oxygen
is supplied to waters through the process of photosynthesis and through diffusion from the
atmosphere. Oxygen is consumed by the respiration of organisms, including animals, plants, and
the decomposing activity of microorganisms. Eutrophication greatly increases the chances of
anoxia and hypoxia by increasing the rate of respiration (Officer el aL 1984; Larsson el aL 1985;
Jensen el aL 1990; Rydberg el aL 1990; EPA 1991; Parker and O'Reilly 1991; Lein and Ivanov
1992). Photosynthesis by phytoplankton produces oxygen, but much of the photosynthesis in
eutrophic waters occurs near the surface, and oxygen readily diffuses to the atmosphere. The
majority of the phytoplankton material is decomposed deeper in the water column, consuming
oxygen there.
Many estuaries and coastal seas are stratified due to density differences resulting from
freshwater running out over denser seawater. Such stratification increases the likelihood of
anoxia and hypoxia, since particulate organic matter sinks into the deeper water but oxygen must
mix down through the pycnocline. However, even in the absence of stratification, eutrophication
can lead to anoxia and hypoxia, as indicated by nutrient enrichment experiment at the Marine
Ecosystem Research Laboratory (MERL) facility at the University of Rhode Island. MERL
consists of a series of mesocosms, large fiberglass tanks containing water and bottom sediments
from Narragansett Bay, designed to mimic the functioning of estuarine ecosystems. In a nutrient
enrichment experiment in which the tanks were kept well mixed, moderate nutrient inputs caused
hypoxia, and anoxia resulted from high nutrient inputs (Oviatt el aL 1986).
Dieback of Seagrasses and Algal Beds
In addition to anoxia and hypoxia, eutrophication can lead to the die-back of seagrass
beds, important habitat and nursery grounds for a variety of fish and other animals. One
mechanism for such die-back is shading out of the grasses by the abundant phytoplankton in the
overlying water, a process thought to have caused the die-back of macrophytes in the upper
portions of Chesapeake Bay (Kemp el aL 1983; Twilley el aL 1985; D'Elia 1987); in the Dutch
Wadden Sea (Gieson el aL 1990), and of both tropical and temperate seagrasses in Australia
(Kirkman 1976; Cambridge and McComb 1984; Cambridge el aL 1986). Die-back caused by
such shading usually manifests itself in a rather gradual loss of the seagrasses (Robblee el aL
1991), although the occurrence of unusual nuisance algal blooms hi 1985 and 1986 greatly
reduced the abundance of seagrass beds near Long Island (Dennison el aL 1989). Nitrogen
96
-------
enrichment may also have a direct physiological response on seagrasses, with internal nutrient
imbalances appearing to lead to reduced survival (Burkholder el aL 1992b).
Beds of attached macro-algae on bottom sediments or rocks can also be adversely
affected by eutrophication. Nutrient enrichment of rocky intertidal areas typically leads to a
reduction hi the overall diversity of both attached algae (Borowitzka 1972; Littler and Murray
1978) and associated annuals (Gappa el aL 1990). These nutrient-enriched areas tend to be
dominated by opportunistic algae with rapid growth rates, such as Cladophora sp. and
Enteromorpha sp. which can take advantage of the elevated nutrient levels and shade out other
species (Littler and Murray 1975, 1978). This is clearly seen along the Swedish coast of the
Baltic Sea, where, since the mid-1970's, nuisance forms of filamentous algae (Cladophora and
Enteromorpha species) have become more dominant, coinciding with a decline of the former
dominant bladderwrack algae, Fucus sp. (Baden el aL 1990; Rosenberg el aL 1990). The
bladderwrack is used as spawning grounds for herring, and the change hi dominance by
macroalgae has led to decreased hatching of herring eggs (Rosenberg el aL 1990).
Nuisance Algal Blooms
Blooms of nuisance algae are characterized by very high abundances in the phytoplankton
of one overwhelmingly dominant species. These blooms often result in noticeable color and are
popularly named by this color: red tides, green tides, brown tides. As with eutrophication
generally, these blooms can result hi anoxic or hypoxic conditions. In addition many nuisance
blooms produce substances toxic to aquatic organisms or humans (Cosper 1991). Green tides
during the 1950's heavily damaged oyster populations on Long Island (Ryther 1954, 1989), and
brown tides hi 1985 and 1986 greatly reduced populations of bay scallops on Long Island
(Cosper el aL 1987; Bricelj and Kuenstner 1989) and of blue mussels hi Narragansett Bay
(Tracey el aL 1989). These shellfish starved to death, since they were unable to graze on the
brown-tide algae. Blooms of some dinoflagellates (red tides) can result hi the accumulation of
toxins in shellfish, which, when eaten by humans, cause paralytic or diarrhetic shellfish poisoning
(Smayda 1989). Frequent blooms of a gold-brown dinoflagellate in Northern Europe have
caused extensive fish mortality since the mid 1960's (Smayda 1989). In 1991, toxins produced
by a diatom bloom concentrated hi anchovy and caused the death of pelicans which fed on these
fish (Work el aL in press, as cited hi Smayda 1992). Production of toxins by diatoms was
completely unknown before 1987 (Smayda 1992). Recently, Burkholder el aL (1992a) discovered
a new toxic dinoflagellate which releases toxins only in the presence of fish and appears to be
responsible for several fish kills in estuaries in North Carolina.
Nuisance-bloom tides have been known since biblical times (Cosper 1991), but blooms
of many species appear to be occurring with greater frequency throughout the world (Hallegraeff
el aL 1988; Anderson 1989; Smayda 1989, 1992; Robineau el aL 1991). Red-tide blooms of
toxic dinoflagellates appear to be more frequent in many parts of the world (Anderson 1989;
Smayda 1989; Wells el aL. 1991), and blooms of cyanobacteria have become more prevalent hi
the less saline portions of Chesapeake Bay (D'Elia 1987) and hi the Baltic Sea and related waters
over the past 10 to -20 years (Smayda 1989 and references therein). Many of the new toxic
97
-------
phytoplankton blooms are sub-populations of previously non-toxic species which now occur at
previously unseen abundances (Smayda 1989, 1992). Brown-tide blooms of Aureococcus
anophagefferens were unknown before 1985 (Sieburth el aL 1988).
The cause(s) of increased nuisance blooms is/are not known, but evidence points toward
the importance of increased nutrient inputs to estuaries and coastal seas. Smayda (1989) has
compiled extensive evidence in support of the hypothesis that the worldwide increase in nuisance
algal blooms is related to increased nutrient availability. For instance, a 2.5-fold increase in
nutrient loadings accompanied an 8-fold increase hi the annual number of red-tide blooms in
a harbor in Hong Kong between 1976 and 1986. Increased nutrient concentrations hi the North
Sea, the Baltic Sea, and in waters between Denmark and Sweden (the Skagerrak and Kattegat)
have co-occurred with increased primary production and increased incidence of blooms hi these
waters (Smayda 1989). The green-tides which occurred hi the Great South Bay of Long Island
in the 1950's were also clearly associated with nitrogen loading from duck farms there (Ryther
1954), and the reduction of nutrient loadings and opening of a channel to increase water
exchange between the bay and ocean have greatly reduced these blooms (Ryther 1989). Also,
nuisance algal blooms are much more likely to occur hi nutrient-rich estuarine waters than in
more coastal or shelf waters (Cosper 1991; Prego 1992).
On the other hand, there is little if any evidence to show a direct connection between
either nitrogen or phosphorus concentrations and blooms of most brown-tide or red-tide
organisms (Cosper 1991; Wells el aL 1991). Red-tide blooms hi Florida are not correlated with
concentrations of any measured form of nitrogen or phosphorus (Rounsefell and Dragovich 1966).
Similarly, the brown-tide blooms of the mid-1980's along the northeastern coast of the U.S. did
not appear to be correlated with higher levels of nitrogen or phosphorus (Cosper el aJL 1989;
Cosper 1991). However, it is important to note that the concentration of a nutrient at any given
point of tune may not be correlated with its availability to phytoplankton (Howarth 1988), and
phytoplankton can grow for long periods of tune off of internally stored pools of nutrients
(Andersen el aL 1991).
Perhaps more importantly, it may not be the availability of nitrogen alone that matters in
controlling nuisance algal blooms, but rather the relative availability of nitrogen hi comparison
to silicon (Officer and Ryther 1980; Smayda 1989). When Si:N ratios are relatively high, silicon
is relatively available, favoring the growth of diatoms, which have a high requirement for silicon.
However, as the Si:N ratio decreases, competition begins to favor other algae with no silicon
requirement, such as the red-tide, green-tide, and brown-tide organisms. Eutrophication can
decrease the abundance of silicon by increasing sedimentation of phytoplankton, as has been
demonstrated hi the Baltic Sea (Wulff el aL 1990). Where long-term nutrient data are available,
the increased occurrence of nuisance algal blooms has always been found to be correlated with
a decrease hi Si:N ratios (Smayda 1989 and references therein). Net primary production probably
remains controlled by nitrogen or phosphorus availability throughout the range of silicon
availabilities (Howarth 1988), but the relative availability of silicon may well control the
abundance of diatoms vs. other phytoplankton species, thereby setting the stage for nuisance
blooms (Smayda 1989).
98
-------
2.2.6.2 Multiple Effects of a Single Class of Contaminants, PCDDs
Not only do the compounds of concern, as a group, generate all of the effects discussed
above, but an individual compound or class of compounds may do so as well. This section
discusses the multiple effects of 2,3,7,8-tetrachlorodibenzo-/j-dioxin (TCDD), considered to be
the most toxic of 75 congeners, or isomorphic shapes, that compose the class of contaminants
polychlorinated dibenzo-p-dioxins (PCDDs). TCDD was the primary source of public health
concern at Love Canal, New York; Seveso, Italy; and Times Beach, Missouri. In different
species, and in different tissues within a species, TCDD is known to cause cancer, impair the
immune system, initiate wasting syndrome, adversely affect the nervous system and behavioral
patterns of individuals, disrupt the endocrine system, and elicit embryo- and fetoxicity, as well
as other reproductive effects, and for laboratory rats and chimpanzees, have transgenerational
effects. That TCDD is responsible for this "perplexing web of interaction" has been explained
by two mechanisms: one which spurs some cell types to grow wildly, and another which inhibits
or causes deviations in some cell types as they differentiate to their respective specialized
functions (Schmidt 1992). Consequently, TCDD has recently been characterized as an
"environmental hormone", because it can alter the functional activity or the structure of various
organs in numerous species. This case study will focus on known human health effects and
implications of the results from laboratory and wildlife population studies.
TCDD is presently considered less of a human cancer risk than was once believed.
However, two recent epidemiological studies support the hypothesis that, at least at relatively
high doses, TCDD can be a human carcinogen. Fingerhut el aL (1991) found that 5,172 workers
from a dozen chemical plants at which exposure occurred had a 15 percent increased chance of
dying from cancer, in comparison to the general population. These findings were based upon
blood serum concentrations of TCDD in 259 of these workers.
Workers with twenty or more years exposure, (including a period in which TCDD
exposure levels would have been higher) exhibited a nine-fold increase in soft-tissue sarcomas
as compared with the general population. Similarly, researchers found a 24 percent higher rate
of death from all cancers in 1,583 pesticide plant workers in Germany, and a 87 percent increase
for a twenty year exposure group (Manz ei aL 1991). Unlike the U.S. study, the German study
did not find an association with any single form of cancer. A critical review of the literature
concluded that because of the array of compounds (including pesticides) also present during any
occupational exposure to TCDD, particularly spraying or other application jobs, it is not yet
possible to assign a causative effect to TCDD alone for malignant lymphomas, and possibly not
for soft-tissue sarcomas (Johnson 1992). This author did find that respiratory system and thyroid
cancers occured at an excessive rate suggestive of a causative role for TCDD.
Both the humoral-mediated immune response, e.g., antibody reactions, and the cell-
mediated immune response, e.g., lymphocyte rejection of foreign tissues or tumors, are affected
in most species (WHO 1989). Recent research (House el aL 1990) has indicated that there is "a
profound suppression of antibody production" in mice exposed to TCDD which occurs in a dose-
dependent fashion, with a significance level less than 0.01. These findings support the results
99
-------
from earlier research (Vecchi el aL 1980; Holsapple el aL 1984). In addition, these authors
suggest that TCDD selectively induces toxicity at the cellular level, thus allowing for multiple
assaults on the host's immune functions. The thymus, and particularly its epithelial cells, are
sensitive to TCDD exposure, as indicated by the occurrence of lesions at levels well below those
inducing lesions in other organs in studies conducted on rats, mice, guinea pigs, and monkeys.
Interestingly, the effect of TCDD on lymphoid tissues is the same hi all species and exerts its
most profound and persistent effects when introduced during the perinatal period (WHO 1989).
Nonetheless, researchers have not found consistent results implicating immunosuppression hi
accidentally exposed humans (Hoffman el aL 1986); however, their offspring have not been
investigated.
TCDD has been found to cause a starvation-like wasting syndrome in all animal species
subjected to acute lethal doses (EPA 1985; Bestervilt el aL 1991). Early studies suggested that
food consumption was decreased, but the reduction of intake could not fully account for the
weight loss (Allen el aL 1975, 1977; Greig el aL 1973; Kociba el aL 1976). Subsequent studies
directed toward the digestive tract could not elicit a generalized impairment of intestinal
absorption (Madge 1977; Manis and Kim 1979; Ball and Chabra 1981; Shoaf and Schiller 1981;
Schiller el aL 1982). Keesey el aL (1976) suggested that body weight in rats is regulated around
an internal setpoint, which is lowered by TCDD. These and other authors found that TCDD-
treated rats vigorously maintained the new, lower setpoint, whether starved or overfed, with the
same precision as the control group (Keesey el aL 1976; Peterson el aL 1984). Wasting
syndrome has been listed as a symptom, although not necessarily confirmed as an effect, of
human exposure to TCDD (ATSDR 1989).
There are a variety of human neurological and behavioral impairments that have been
associated with acute exposure to TCDD or mixtures containing TCDD, including sexual
dysfunction (lack of libido and impotence); headache; abnormal nerve conduction and clinically
uncorroborated joint pains; sleep disturbance; depression; loss of energy and drive;
uncharacteristic bouts of anger; and possibly sight disturbance and loss of hearing, taste, and
smell (Fillipini el aL 1981; WHO 1987). There have been only two cases of exposure to "pure"
TCDD, which involved a total of seven people. The exposed class, as a whole or individually,
exhibited all of the above symptoms, sometimes for up to two years after exposure. There were
also individual instances of hirsutism, chloracne, and other effects indicating alterations in body
chemistry. It was considered likely, but not conclusive, that the delayed manifestation of these
symptoms was due to the original TCDD exposure.
Human health effects at the individual and population level from chronic exposure to
TCDD have not been identified. However, a critical need hi future research can be identified by
examining the results from experimental studies and research on wildlife populations with regard
to behavioral impairments, endocrine system alterations, reproductive and developmental toxicity,
and transgenerational effects.
A subtle form of behavioral impairment has been identified hi a multigenerational
experiment involving non-human primates. Schantz and Bowman (1989) found a dose-
100
-------
dependent relationship in the offspring of female rhesus monkeys which were fed daily diets
containing 0 ppt, 5 ppt, and 25 ppt of TCDD. Several years aftei secondary offspring exposure
(in utero and four months of nursing) had ceased, these authors found a dose-dependent
relationship for spatial discrimination reversal learning (DHL) and suggested a NOAEL of 5 ppt.
Bowman el aL (1989) expressed concern that this may be an artificial NOAEL because the
TCDD lipid values were assumed to be zero for the offspring of the controls, which actually may
have background concentrations of TCDD-like substances, such as furans and PCBs, that could
elicit the same effects, and because individuals varied greatly in their abilities to metabolize the
dose received from the mother. Similar infant exposures to PCBs have been correlated with
subtle cognitive impairments (Rogan el aL 1988; Swain 1988; Jacobson el aL 1990; Tilson el aL
1990; Jacobson el aL 1992). The ultimate impact of these individual cognitive impairments can
be characterized as a "diminishment of potential" in humans.
Endocrine disruption, reproductive and developmental effects, and transgenerational effects
have distinct profiles resulting from acute doses, but the distinctions blur somewhat when
considering lesser exposures. TCDD exerts antiestrogenic, estrogenic, and antiandrogenic effects
on the endocrine system resulting in inter alia, decreased uterine weight, estrogen-induced protein
secretion, and estrogen and progesterone receptors; and decreased androgen secretion, reduced
testosterone levels by 90 percent, testicular atrophy, reduced fertility, and decreased
spermatogenesis (See Effects on Reproduction). Reproductive effects include morphological
changes in the ovaries and uterus of rats (Kociba el aL 1976), reduced conception rates and a
high incidence of early spontaneous abortions in monkeys (Allen el aL 1977; Barsotti el aL
1979). Peterson el aL (1992) have found an EDjo of 0.16 ppb in rats, based on a single maternal
dose on Day 15 of gestation. Peterson found indications of demasculinization at the lowest dose
administered, 0.064 ug/kg body (64 ppt). He has not determined a NOAEL. This dosage was
transferred to the pups in utero and through lactation, to be associated with a range of adverse
effects in the development of the male reproductive system and in behavior, including delayed
and incomplete organ development, inhibition of spermatogenesis, both demasculinization and
feminization .of sexual behavior, and alteration of the regulation of the luteinizing hormone.
Lowered sperm production of 75 percent did not affect the rats' fertility. Normally, rats ejaculate
up to ten times the amount of sperm needed to ensure pregnancy.
Developmental toxicity can be described in terms of embryo/fetotoxicity, structural
malformations, and postnatal functional alterations (USEPA Draft 1991). Except for the hamster,
the lethal effect of TCDD on the fetus is likely secondary to maternal toxicity, i.e., the fetus dies
only when there are apparent adverse effects on the mother from the dose. Structural
malformations include thymic hypoplasia, hematological alterations, subcutaneous edema, extra
ribs (rabbit), cleft palate malformation (mouse), and intestinal hemorrhage (rat). There have been
two studies focusing solely on the transgenerational effects of TCDD. One involves the effects
of exposure on the reproductive system and behavior of .rats (Murray el aL 1979), and the other
on the reproductivity and behavioral effects on rhesus monkeys (Bowman el aL 1989). Murray
el aL (1979) conducted a three generation reproductive study on Sprague-Dawley rats fed daily
diets containing 0, 0.001 ppm, 0.01 ppm, or 0.1 ppm TCDD. The groups in the first generation
were fed for 90 days prior to mating. No effect on mating frequency was observed, nor were
101
-------
any toxic effects. However, the offspring and third generation that were then also fed a diet
containing 0.01 ppm TCDD per day showed decreased body weight and reduced food
consumption. The first generation's fertility was greatly reduced at a dosage of 0.01 ppm per
day, and the second and third generations' fertility levels were significantly reduced at dosages
of 0.001 and 0.01 ppm per day, respectively. The 0.01 ppm dosage also resulted in reduced litter
sizes, an increase hi feto- and neonatal mortality, and a decrease in postnatal growth. As a
result, 0.001 ppm per day TCDD was suggested as a NOAEL for reproductive lesions. However,
revaluation of the same data from a transgenerational perspective (all generations statistically
pooled) indicated that 0.001 ppm did have a statistically significant effect, and thus should not
be used as a NOAEL (Nisbet and Paxton 1982). This level of effect is supported by additional
reevaluation of these data by Allen el aL (1989) and by data from the rhesus monkey study
(Schantz el aL 1989).
The potential human health impact of TCDD exposure based on the sum of known
endocrine, reproductive, and transgenerational effects in experimental and wildlife populations
includes: (1) TCDD has an extended half-life and can thus keep a gene "on" or "off for an
excessive amount of tune, or be transferred in utero or through lactation to the next generation
in sufficient amounts to cause harm. Because of this extended biological half-life and the
apparent absence of a threshold for adverse effects, the reproductive system appears to be the
most sensitive to TCDD exposure, particularly during the perinatal period; (2) there is existing
evidence which suggests that prenatal androgenization affects human sexual behavior and
structure of the hypothalamus (Erhardt and Meyer-Bahlburg 1981; Hines 1982; LeVay 1991),
thus altering the nature of human reproductivity; and (3) unlike rats who ejaculate 10 tunes more
sperm than needed for successful fertilization, humans have almost no margin for error in terms
of successful insemination (Carlsen el aL 1992). Consequently, impairment of spermatogenesis
would likely have a negative impact on human fertility (Peterson el aL 1992). Thus, it is
possible, but not yet demonstrated, that the cumulative impact of chronic and in utero exposures
humans receive have been and/or are affecting both the nature and success of human
reproductivity at the population level.
2.2.6.3 Effects Of Multiple Compounds of Concern
On a Single Species: Forster's Tern
A case study of the Great Lakes Forster's tern provides an example of the difficulty in
recognizing subtle effects and sensitive endpoints resulting from ambient exposure to multiple
chemicals over time. Overt endpoints of high-dose exposure, such as birth defects and outright
mortality, are far easier to observe than low-dose functional deficits that are not expressed
immediately after birth. Consequently, as conditions of the environment improve and exposure
levels decrease, less visible, widespread health decrements in wildlife and human populations
could be missed as the following case study demonstrates.
A cross-disciplinary team of researchers observed a colony of troubled Forster's terns
(Sterna forsten) in Green Bay in 1983 and 1988 (Hoffman el aL 1987; Kubiak el aL 1989). The
102
-------
study population was a colony of nesting Forster's tems on a confined waste disposal facility in
Green Bay, Lake Michigan, Wisconsin. The tern control population was nesting on an inland
lake and not dependent upon food sources from the Great Lakes. Nesting success was recorded
and samples of eggs and chicks were collected for chemical and in vitro analysis of
bioaccumulative contaminants. In 1983, tern offspring experienced significantly poor hatchability
(37 percent compared with controls at 75 percent), low chick body weight, increased ratio of liver
to body weight, edema, reduced fledgling success, and lack of parental care compared with the
in-land population (Kubiak el aL 1989). Seventeen days after hatching, 35 percent of the chicks
had died. In one component of this study, an egg exchange experiment among the Green Bay
colony, the control colony, and laboratory incubators revealed that embryotoxicity, chick
mortality, and parental abandonment contributed to the lack of nesting success of the Green Bay
tems.
Significantly higher concentrations of PCBs and dioxins were found hi the Green Bay
colony. Tissue culture bioassay for AHH enzyme induction revealed significantly higher enzyme
activity measured as dioxin toxicity equivalents (TEQs) in the Green Bay population than
controls. Going one step further, this was confirmed using PCB congener-specific chemical
analysis and multiplying AHH enzyme induction toxicity factors by the quantities of specific
congeners in chicks and abandoned eggs. The congener-specific chemical determination revealed
that 95 percent of the toxicity was from PCBs and about 3 percent from dioxins.
The scenario at the Green Bay colony changed considerably in 1988 (Harris 1990)
although the final outcome was similar. The median total PCBs hi the eggs hi 1983 was 22.2
ppm. In 1988, the eggs held 7.3 ppm (median), a 66 percent reduction. Dioxin enzyme
induction toxicity equivalents declined 58 percent, from 2175 to 913 (201 enzyme-induction
TEQs in the referent population). Certain endpoints — hatchability, length of incubation, weight
gain, and number of young fledged — were normal and did not deviate significantly from the
1983 control population up to 17 days posthatching. However, hi the latter quarter of
development, commencing on day 18, the chicks showed signs of wasting and by day 31, 35
percent of the young had died. This was the same proportion that had died in 1988, but two
weeks later. Thus far, wasting appears to be the most sensitive endpoint researchers have
identified hi Forster's terns as a result of exposure to dioxin-like contaminants. If the higher-
dose endpoint of hatchability, an obvious and easy endpoint to measure, had been used as the
only endpoint of the second study, the delayed, but equally devastating effect of wasting would
have been missed.
Other latent effects hi the Forster's terns were not reported because the short-term and
long-term fate of the chicks that fledged was not determined beyond day 31. Long-term
banding and breeding population assessments have not been conducted to determine if this
population of Forster's tern existed because of immigration of breeding birds from clean areas
as is the case with Great Lakes bald eagles (Hqliaetus leucocephalus) and Caspian terns
(Hydroprogne caspia) (Colborn 1991, L'Arrivee and Blokpoel 1988).
103
-------
Two facts are worth noting: (1) no Forster's terns have returned to the Green Bay Island
since 1988 (Ludwig 1992); and (2) no lesion for wasting has ever been identified. A laboratory
study in which 2,3,7,8-TCDD was administered to rats intracerebroventricularly into the lateral
brain ventrical and subcutaneously at the back of the neck at a pumping rate of 1 u,l/h or 20-21
Hg/kg body weight per day induced wasting only in the brain treated animals, suggesting that
wasting may be the result of central nervous system damage (Pohjanvirta el aL 1989).
22.1 Conclusion
Atmospherically transported toxic contaminants impacting the world's great waters
represents one of the largest challenges facing the scientific and managerial communities today.
The problems associated with identifying and ultimately managing the sources, fate, transport,
effects, control, and remediation of toxic contaminants in large marine and aquatic ecosystems
are among the most difficult contemporary issues confronting environmental managers and
decision-makers.
While loadings and inputs of toxic chemicals are direct, variable, and waterbody specific,
it is clear that all of the world's great waters are being perturbed by contributions of toxic
substances from the atmosphere. In most cases, the sources driving the atmospheric
concentrations are poorly understood, and the dimensions of the airsheds for each of the world's
great waters are largely unknown. An increasing body of evidence indicates that long-range
transport of atmospheric contaminants results in transboundary pollution of the world's great
waters, and that this mechanism does not respect geographical, political, jurisdictional, or national
boundaries.
The fate of toxic substances hi large marine or aquatic ecosystems is presently
incompletely understood, but it is recognized as critically important because of the uptake of
contaminants by native biota. Within the waterbody, the phenomenon of biomagnification often
results in excessive increases in contaminant concentration at each succeeding trophic level in
the food chain. Food chain accumulation ultimately leads to human exposure, as humans are one
of the final predators hi the great waters ecosystems.
The data presented in this report repeatedly demonstrate that all of the ecosystem
compartments of the world's great waters — i.e., the atmosphere, the water column, the
sediments, and the biota, including humans — are irrevocably interrelated, interconnected, and
reciprocally interactive. They further indicate that by the tune the sources, fate, transport, and
effects of a toxic compound are identified and understood, it is too late, and the inevitable
impacts of those materials on the system will have occurred. Therefore, in addition to
remediating past inputs, a philosophy of prevention is mandated. In order to respond to this
challenge, the regulatory community will be required to implement a prevention policy which is
guided by a perspective of our interrelated environment, and which extends beyond both
environmental compartments, and local, state, provincial, regional, national, and international
boundaries.
104
-------
Overview of the Current State of the Great Waters.
As a result of our increased understanding of the effects of nutrient additions and the
implementation of control practices, eutrophication is beginning to be managed in many of the
world's great waters. For a number of these systems, water clarity has improved and anoxia has
been minimized. While significant improvement has been made for many of the great waters in
the last two decades, some areas still require additional efforts.
Toxic residues in some of the ecosystem compartments of many of the world's great
waters have begun to decline. However, the observed rates of decline have recently decreased,
and, in many areas, it is considered inadvisable to consume the biota of these waters. In many
of these systems, obligate fish consuming wildlife are adversely impacted, and frequently fish
stocking is required because of reproductive failures in the fish populations. In many areas, fish
consumption advisories are in effect as a part of an effort to minimize or eliminate negative
impacts of toxic chemicals on human health. The slow response times of many of these bodies
of water suggest that extended periods of time, on the order of decades, will be required before
these systems recover completely from past and present chemical insult, even when all sources
of toxic substances are eliminated.
In summary, for most of the great waters, present conditions are significantly improved
as compared with two to three decades ago. However, the majority, if not all, of these systems
are far from fully recovered.
Chemical Contaminant Profile Summaries.
This section summarizes the present state of knowledge and the current status for a
number of compounds known to be atmospherically transported to the world's great waters. Each
major chemical or contaminant class of compounds is considered individually below.
2t3>7,8~Tetrachloro-p-dibenzodioxin.
As long as industrial society continues to depend upon incineration and combustion
processes as a source of energy, a means of waste disposal and a process of production, TCDD
will be a source of concern. Present concentrations of 2,3,7,8-TCDD in human adipose tissue
are globally quite consistent in the 5 to 10 ppt range. However, because the analytical techniques
required to measure dioxins have only recently become standardized, there is no present method
available to estimate whether body burdens in the human population are increasing or decreasing
as compared with historic backgrounds. The non-carcinogenic effects of dioxin have recently
received increasing attention, and appear to be as subtle, and possibly more serious, than the
potential for cancer. Dioxin is still considered the most toxic xenobiotic substance produced by
human activity. While its effects are dramatically different among various species, the greatest
exposure pathway in most instances is the ingestion of contaminated foodstuffs. Fetuses and
nursing infants are at exceptional risk to exposure, even more so than individuals eating 2,3,7,8-
TCDD contaminated fish.
105
-------
Cadmium
Cadmium exposure is an excellent example illustrating the fact that a relatively constant
low-dose exposure from multiple pathways can produce a slow, but steady, increase in the body
burden of the contaminant in a population. Worldwide body burdens of cadmium are rapidly
approaching the maximum safe tolerance limits. Inhalation of cigarette smoke is the most
important exposure pathway, with consumption of contaminated foodstuffs a close second. Gross
teratological and behavioral changes have been reported in experimental animals following
cadmium exposure. Low birth weight has been associated with cadmium exposure in both
animals and humans. Long— term industrial exposure to cadmium has been reported.
Even though production of chlordane for domestic use has ceased in the United States,
commercial products containing this pesticide are still available until the stocks are depleted.
Chlordane and its metabolites in fish have been associated with areas of urbanization, suggesting
its misapplication, possibly against termites. In the Great Lakes, oxychlordane concentrations
in fish tissue are regarded as having reached a level of concern.
The principal exposure pathway is generally food. However, both inhalation in homes
treated with chlordane, or ingestion of contaminated drinking water could become primary
pathways hi areas where this pesticide was used or disposed of carelessly. An association
between fish consumption and human residues of chlordane metabolites has been reported.
Chlordane both induces enzyme production and disrupts endocrine control.
DDT/DDE
Concentrations of DDT in human tissue are decreasing; however, its biodegradation
product, DDE, does not appear to be declining. Since DDT is not readily converted to DDE in
humans, and human residues are declining, it is assumed that the food pathway is contributing
to present body burdens of DDE. Although its use has been banned in Canada and the United
States, long-range transport of DDE to the great waters will be a continuing problem. DDE is
an enzyme inducer, gap junction intercellular communication blocker, and disrupts endocrine
control. Concentrations in maternal breast milk have been associated with hyporeflexia in
neonates. Human tissue levels of DDE have been associated with the consumption of fish.
Dieldrin
Although the manufacturing and large number of uses of dieldrin have been banned in
the U.S., there does not appear to be a decline in human residue levels to date. Dieldrin
accumulates in human tissue with age and is preferentially transferred to the fetus via the
placenta and to the newborn in breast milk. This toxic substance is an enzyme inducer, gap
junctional intercellular communication blocker, and disrupts endocrine hormone control.
Exposure likely results from leaching of residuals from past use and improper disposal.
106
-------
Hexachlorobenzene (HCB)
Hexachlorobenzene is created unintentionally during the production of pesticides and the
combustion of chlorine containing material. As a result, it is ubiquitous in the environment.
Tissue residue surveys find that HCB concentrations have not declined since 1975 and suggest
that concentrations may be increasing. However, food residues in some highly contaminated
areas of the U.S. have shown a decline. HCB is capable of enzyme induction and disruption of
endocrine control. Severe, long-lasting health effects have been seen in a cohort of people
exposed to high concentrations of HCB after eating HCB-treated seed; 95 percent of all in utero
infants at the time of the incident died within two years of birth. There were many stillbirths as
well. Nursing infants ingest 200 to 300 times the adult intake on a bodyweight basis.
Significantly higher concentrations were found hi cadavers from Kingston, Ontario when
compared with Ottawa, Canada. Similarly elevated concentrations of HCB were found in
follicular fluids in persons living near Hamilton Harbor when compared with those from other
southern Ontario communities. In the Great Lakes, HCB concentrations in fish and water were
reported at a level of concern in 1986.
Lead
Recent efforts in lead research have revealed new subtle health effects not previously
recognized. These observed impacts included neurological, immunological, developmental, and
reproductive effects. Maternal prenatal exposure has been associated with low birth weight,
shortened gestational age, neurobehavioral, and psychomotor deficits in offspring, confirming that
lead is a human neuroteratological agent.
Strong associations have been found between lead exposure and detrimental effects on
behavior, cognitive, and motor development of infants and children. Because its
immunosuppressive actions have been demonstrated in laboratory animals at very low doses, the
potential for effects in humans merits serious consideration.
Lindane (Isomer of Hexachlorocyclohexane; HCH)
Isomers of HCH do not appear to be decreasing in human tissues. The alpha isomer of
HCH was established to be at a level of concern in the Great Lakes hi 1986. The estrogen
effects of lindane and its adverse effects upon the male reproductive system have been reported
in a variety of animal studies. Because of human breast milk concentrations of this pesticide,
nursing infants are at special risk. Lindane induces enzymes, blocks intercellular gap junction
communication, and interferes with endocrine control.
Mercury
Human exposure to mercury is associated with both naturally contaminated bodies of
water and marine and freshwater ecosystems hi which mercury has accumulated as a result of
industrial activity. Methyl mercury is of special concern because it is completely absorbed upon
107
-------
ingestion. Under anaerobic conditions in lake sediments it is converted from metallic mercury
to the methyl form and readily bioaccumulates in fish tissue. A number of studies have shown
a correlation with human mercury residues and fish consumption. An association with the
number of dental fillings of mercury amalgams and mercury residues in blood and urine has been
reported. In animals, methyl mercury preferentially crosses the placental barrier and the fetal
blood brain barrier, and is neuroteratological.
Polynuclear Aromatic Hydrocarbons (PAHs)
If estimates of continued fossil fuel combustion are realistic, PAHs are going to be a
continuing problem for the world's great waters. Improvements in analytical technology have
revealed that PAHs bioconcentrate in certain tissues, which was not considered possible in the
past because of their rapid enzyme induction capacity. There is no information available to
predict the human health effects of PAHs. PAHs tend to accumulate in the sediments associated
with great waters, and have been implicated in a variety of tumors and cancers associated with
bottom-dwelling fish. Many of the PAHs are potent carcinogens, and some have been shown
to be genotoxic agents.
Polychlorinated Biphenyls (PCBs)
Although PCB production has been banned North America, it is estimated that more than
50 percent of total production is still in use. Because of this enormous reservoir, the persistence
of this group of compounds, and inadequate disposal methodologies, PCBs will likely continue
to be a major problem in the world's great waters. Although pathways contributing to
background human exposure have not been clearly defined, a number of studies suggest that
inhalation is a minor pathway. Several of the tetra-, penta-, and hexachlorobiphenyls are known
inducers of AHH/EROD enzymes, and have been associated with thymic involution,
teratogenicity, "wasting", and porphyria hi a number of laboratory animals. Some PCB congeners
are more toxic than others. These forms induce enzymes, block intercellular communication, and
disrupt glucocorticoid control. They have been associated with developmental decrements and
reduced birth weights in human infants and with shortened gestation periods. It has been
suggested that as PCBs recycle in the world's great waters, the more highly chlorinated
(potentially more toxic) congeners will become a larger component of the total PCB
concentration in circulation.
Toxaphene
The pesticide toxaphene is a mixture of 177 compounds about which little is known. Its
use has been limited. Because of its persistence, biomagnification and dispersal potential via
long-range transport, it will continue to be of concern in the world's great waters. In very high
doses compared to ambient concentrations, it has been found to be an enzyme inducer, gap
junction intercellular communication blocker, and interferes with endocrine control. Toxaphene
is listed by USEPA as a Class B2 carcinogen.
108
-------
22.9 Application of New Knowledge
Related to Toxic Substances
One of the major needs relative to airborne toxic substances is a methodology which will
reliably express the biological toxicity or potency of these compounds. With this tool in hand,
a method for quantification of impacts and effects against relative toxicity would be available.
This is particularly important when groups of compounds such as polychlorinated biphenyls
(PCBs), polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs)
are considered.
The PCS group of compounds consists of 209 theoretically possible congeners, the PCDD
group of substances are comprised of 75 congeners, and the PCDF group of compounds consists
of 135 congeners. Each of these congeners are related to the original parent compound, but each
differs slightly in degree and position of chlorination, in stereochemistry, and, most importantly,
in biological toxicity or potency. The PCB group of compounds probably affords the best
example for consideration.
Early lexicological research treated PCBs as a series of commercial mixtures. Normally,
results were described as "Total PCBs" or as an Aroclor mixture. In either case, the reference
Aroclor was used, ignoring the fact that it consisted of up to 50 or more congeners of PCB, each
with varying toxicity. To date, all of the epidemiological studies performed have relied upon the
use of "Total PCBs" as a measure of toxicity resulting from exposure. However, there is a
growing body of evidence which suggests that only a relatively few highly toxic PCB congeners
may be responsible for many of the observed outcomes of exposure (Jacobson el aL 1989;
Kubiak 1988; Kannan el aL 1988; Bush el aL 1984 and 1985).
These few highly toxic PCB congeners are generally planar or nearly planar in nature.
The planar or nearly planar group of substances include not only non-ortho and mono-ortho
substituted PCBs, but also polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated
dibenzofurans (PCDFs).
Although the various planar congeners of PCBs, PCDFs and PCDDs differ widely in their
biological toxicities, they are all quite similar in their stereochemistry and produce similar,
characteristic patterns of toxic responses in mammals (Poland and Knutson 1982; Safe 1987;
Tillitt el aL 1988a and b). Tillett el aL (1988b) states that it is generally accepted that the toxic
properties of various planar chlorinated hydrocarbon compounds are expressed as a function of
a common mode of action. Given this fact, it is, therefore, possible to calculate the biological
toxicity or potency of any of these compounds either individually or in complex mixtures. This
expression of potency is usually made in relationship to the most toxic of the planar chlorinated
hydrocarbons, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Ample precedent for this
assignment of toxicity as TCDD-equivalents exists (Eadon el aL 1986; Safe 1987; Tillitt el aL
1988a and b; and Kubiak 1988). The usual mechanism employed to evaluate TCDD-equivalent
toxicity is to measure the ability of the individual planar chlorinated hydrocarbon to induce mixed
function oxidase enzymes in cultures of liver tissue cells. These enzyme assays include aryl
109
-------
hydrocarbon hydroxylase (AHH) and the cytochrome P-450-dependent ethoxyresonifin-o-
deethylase (EROD) in rat hepatoma cell cultures. The magnitude of the enzyme response for an
individual planar compound or a complex mixture of these substances is then expressed relative
to the magnitude of the response elicited by the most toxic planar compound, 2,3,7,8-
tetrachlorodibenzo-p-dioxin, as TCDD-Equivalent Toxicity. The estimation of TCDD
Equivalents has been shown to correlate strongly with the observed toxicity in mammals of
various individual compounds and mixtures of PCB, PCDF, and PCDD congeners (Sawyer el aL
1984; Mason el aL 1985; and Safe 1987). TCDD equivalents are also variously referred to as
dioxin equivalents, toxic equivalencies (TEQs), or toxic equivalency factors (TEFs). Authors will
also frequently combine these various designations, e.g., Dioxin-TEFs, TCDD-TEQs.
Kubiak (1988) has prepared a series of conversion factors (Keq) for determining "2,3,7,8-
TCDD Equivalents" for compounds isosteric with 2,3,7,8-TCDD, based upon this relative ability
of the planar substance to induce aryl hydrocarbon hydroxylase (AHH) and ethoxyresorufin-o-
deethylase (EROD). A listing of PCBs and then- associated 2,3,7,8-TCDD equivalents is
presented hi Table 30. In practice, the Keq values are simply multiplied by the concentration of
the individual congener to estimate the toxic equivalency (TEQ) or the toxic equivalency factor
(TEF) relative to 2,3,7,8-ietrachlorodibenzo-p-dioxin.
The EROD and AHH enzyme induction tests provide separate and independent estimates
of the potency or the biological toxicity of a given compound or complex mixture. Under normal
circumstances, the values derived from these tests are in good agreement with each other. Since
the induced enzyme levels correlate strongly with observed toxicity hi mammals, either EROD
or AHH results may be reasonably used to estimate the toxicity of those planar chlorinated
hydrocarbon compounds whose chief mode of action is enzyme induction. In practical
application, a single enzyme induction value is usually derived for either an EROD or AHH
induction test for each congener or complex mixture tested. This value is then used to represent
the potency of that congener or mixture.
By comparison, TCDD-equivalency values for individual congeners are calculated hi the
same fashion (Table 30), but direct evaluation of complex mixtures of chlorinated hydrocarbons
using TCDD-equivalents must be undertaken with some degree of caution. This is necessary
because the calculated TCDD-equivalent value of the sum of the planar compounds will often
exceed the value observed upon testing of the mixture by either EROD or AHH protocols. There
appear to be two likely reasons that the simple sum of the TCDD equivalent values tend to
slightly overestimate the actual enzyme induction observed. While the exact mechanisms are yet
unknown, it is known that the toxic interactions between and among the various planar
compounds have been shown to exhibit synergism, additivity, or antagonism (Bimbaum el aL
1985; Eadon el aL 1986; Keys el aL 1986; Bannister and Safe 1987). Secondly, it appears that
the non-toxic,or relatively non-toxic, non-planar congeners contained in complex mixtures of
compounds also tend to compete for the same substrate binding sites as the planar congeners.
Since fewer binding sites are available for the more toxic planar structures, proportionately less
opportunity exists for induction of enzymes than hi the case of a planar constituent measured
individually (J. Ludwig, personal communication 1992). The importance of this interaction is
apparent when the toxicities of the various Aroclor Standards are compared with the more active
enzyme inducers (Table 30).
110
-------
TABLE 30
CONVERSION FACTORS (KEQ) FOR DETERMINING
"2^,7,8-TCDD EQUIVALENTS" FOR PCB CONGENERS
BASED UPON RELATIVE ABILITY TO INDUCE AHH AND EROD ENZYMES
Compound
2,3,7,8-TCDD
3,3',4,4',5-PeCB
3,3',4,4'-TeCB
2,3,3',4,4'-PeCB
2,3,3',4,4',5'-HxCB
2,3,4,4',5-PeCB
2,3,3',4,4',5-HxCB
2'3,4,4',5-PeCB
2,2',3,3',4,4',5-HpCB
3,4,4',5-TeCB
2,3,4,4'-TeCB
2,3,3',4,4',5,5t-HpCB
2,3',4,4',5-PeCB
2,3',4,4',5,5'-HxCB
2,2',3,4,4'^I-HxCB
2,3,3',4,4',6-HxCB
2,3,4,4',5,6-HxCB
Aroclor 1232
Aroclor 1248
Aroclor 1242
Aroclor 1254
Aroclor 1268
Aroclors 1260 and 1262
Aroclor 1016
Aroclor 1221
AHH ERQD
1.00 1.00
126 0.40 0.32
077 0.0027 0.009
169 0.0016 0.0033
105 0.0011 0.0006
157 . 0.000135 0.000063
114 0.000095 0.000142
156 0.000046 0.000089
123 0.000024 0.000012
170 0.000016 0.0000066
081 0.0000086 0.0000417
060 0.0000085 0.0000417
189 0.0000085 0.0000102
118 0.0000083 0.0000091
167 0.0000072 0.0000089
138 <0.0000072 <0.0000089
158 <0.0000072 <0.0000089
128 <0.0000072 <0.0000089
166 <0.0000072 <0.0000089
0.0019394 0.0000019
0.0000173 0.0000163
0.0000137 0.0000185
0.0000099 0.0000131
0.0000057 0.0000051
Active inducers, not quantified.
No induction.
No induction.
Source: Kubiak (1988)
111
-------
Based upon the information provided by extensive testing in wildlife populations and
limited application to human health considerations, it would appear that the use of congener-
specific analysis would offer far more specificity and enhanced resolution in research related to
the effects of toxic substances. The idea of equating of the degree of toxicity with the quantity
of total PCBs, PCDDs, or PCDFs observed is obviously in error. The availability of new analytic
techniques capable of measuring low levels of these compounds by congener, coupled with AHH
and EROD enzyme induction assays, offer the potential to consider observed investigative
outcomes in the light of more reliable toxicity data using dioxin equivalents.
Ultimately for wide application of these techniques, it will be necessary to alter the
regulatory requirements for analytical testing to include congener-specific methodologies, rather
than the existing comparisons with Aroclor standards.
22.9 Future Research Needs
2.2.9.1 Introduction
It is clear that although progress is being made towards the identification of airborne
water pollutants and understanding their biological effects in wildlife and humans, there remains
much that needs to be done. The mechanisms of action and diversity of effects of most
xenobiotics are still not completely understood. However, the power of basic scientific research
has been demonstrated with the identification of carcinogens and their modes of action.
The dominance of cancer as the effect of primary concern in assessing the risk of
pesticides is being challenged by new evidence of effects of chemicals on the nervous, immune,
endocrine, and reproductive systems of laboratory animals, wildlife, and humans. The disease
state, or effect, hi this case is measured by loss of function rather than gross clinical endpoints.
Furthermore, it is now perceived that functional deficits in humans as a result of exposure to the
chlorinated compounds, PCBs and dioxins, occur at lower concentrations than those extrapolated
in rodent models to cause cancer. Most of the research on developmental toxicity has been done
on PCBs and dioxins and on only a few chlorinated insecticides. As a result, little is known
about the non-cancer health effects of pesticides and especially herbicides, the largest portion
on a weight basis of pesticides currently in use. Of concern, are the infrequent and occasional
studies that have shown without a doubt that many of the widely used pesticides are capable of
interfering with the development and function of one or more of the critical life systems.
Because of the potential threat to wildlife and human populations of these findings it is
imperative to establish the means to better understand the non-cancer health effects of (1) all
pesticides in use, (2) those that have been banned or restricted, and (3) any new pesticides being
registered. To delay could seriously affect the survival and well-being of future generations.
As a result of the great diversity of effects, the complicated mechanisms of action, and the
insidious nature of low-level exposures, increased and broad-based funding for innovative
research on non-carcinogenic end-points and mechanisms in wildlife and humans is clearly
warranted.
112
-------
The following identified research needs are prioritized within general fields of research.
However, the fields themselves are not prioritized, since all fields of research must progress
together to achieve a proper understanding of the problem. These prioritized needs are intended
to identify some of the more apparent gaps in our knowledge in each general field of research.
Obviously, these lists can not be comprehensive, but they will serve as a guide for researchers
and funding agencies alike.
2.2.9.2 Research Needs Related to Eutrophication
1. Atmospheric nitrogen is delivered to coastal waters both through direct deposition to the
waters and through deposition on upstream watersheds followed by gradual downstream
washout. The extent to which nitrogen deposited on watersheds is retained in the
watershed rather than being exported downstream is very poorly known and probably
varies greatly depending upon a variety of factors, including land use in the watershed
and age of forest stands. Research on these factors is required if we are to better
understand the importance of atmospheric nitrogen on coastal eutrophication. Such
research may lead to control strategies beyond simply controlling atmospheric nitrogen
emissions, such as managing forest growth or wetlands which fringe streams.
2. Increased nitrogen inputs are well known to be the dominant cause of eutrophication
(overall increased algal growth, causing anoxia, hypoxia, and dieback of macrophyte beds)
in many, perhaps most, of the estuaries and coastal waters of the United States. However,
it is much less clear that nitrogen is the cause of the increased incidence of nuisance algal
blooms by single species of algae (red tides and brown tides). Research is needed to
determine: 1) if nitrogen alone is a proximate cause of blooms; 2) if eutrophication from
increased nitrogen loading might result in the formation of nuisance algal blooms
indirectly (for example by lowering the availability of silica or by increasing the extent
of anoxic sediments); 3) if some other element such as iron or molybdenum must interact
with nitrogen to trigger a bloom; or 4) if nitrogen has no relationship to bloom formation
in the coastal Great Waters.
3. Most dose-response relationships for nitrogen and coastal eutrophication have dealt with
annual time steps. However, it may be only necessary to control nitrogen deposition
during some critical period of the growing season in some coastal Great Waters. The
seasonal variation in the response of esruarine eutrophication to nitrogen inputs from
atmospheric deposition requires further research. Factors to consider include the spatial
and temporal patterns of nitrogen transport in the atmosphere, the residence time of
nitrogen in watersheds, and the seasonality of phytoplankton production in estuaries.
4. Increased nitrogen inputs to many coastal waters and estuaries leads to increasing
eutrophication and anoxia and hypoxia (low oxygen in the water column). Research is
needed to determine if this increases the sensitivity of the biota to other stresses, such as
those from toxic substances.
113
-------
2.2.9.3 Research Needs for Ecosystem Level Effects of Xenobiotic Substances
Even though studies of the long-range atmospheric transport of toxic xenobiotic chemicals
began as early as the mid-1970s, the scientific community only has a limited understanding of
a variety of issues surrounding the central question. Upon reaching the aquatic or marine
ecosystem, a further array of questions remain unanswered. Research on the spectrum of these
issues is required if understanding of fate and transport of toxic chemicals is to be achieved.
1. Our present knowledge of the rate and magnitude of inputs of toxic substances to the
world's Great Waters is extremely limited.
2. Additional research on sources of these contaminants is required, with special emphasis
on differentiation between such issues as revolatilization, existing domestic sources, and
transboundary pollution from foreign sources.
3. The contemporary understanding of deposition processes is limited. Additional research
on the mechanisms involved in the entry of these compounds into waterbodies is required,
as is study of the form of the materials entering the ecosystem. Recent studies suggest
that some of the assumptions made about deposition processes have been incorrect.
Additional studies are required for verification.
4. The understanding of the scientific community of the bioavailability of these chemicals
is limited. Additional research is required to understand the fate of these compounds and
the ultimate exposure of biota in the Great Waters. This knowledge would resolve the
question of concentrations of chemicals versus the estimates of biota exposure.
5. Research addressing "new, relatively unstudied" contaminants, e.g., atrazine, entering the
ecosystem, is required.
6. Research is needed on the effects of pH, temperature, salinity, and dissolved oxygen on:
(1) the internal response of the organism; and (2) the effective dose to the organism.
7. Research is also needed on determining the assimilation efficiencies for a variety of
chemicals in various organisms.
8. Additional field studies on the effects of these materials, particularly subtle effects, are
required.
9. One of the most promising areas of research includes the integrated study approach
incorporating fate assessment chemists, biologists, and lexicologists. These studies will
assist in establishing and defining cause-effect linkages between airborne toxic
compounds and receptor organism effects.
114
-------
2.2.9.4 Research Needs for Wildlife and Human Health Effects
from Xenobiotic Substances
1. Current research on most of the wildlife health problems and some of the human health
problems induced by xenobiotic contaminants often results from serendipitous
observations by scientists engaged in other field or laboratory studies. In the light of the
present evidence, a new vehicle is needed to enable and encourage forensic research
demonstrating the effects of chemicals in living organisms. The organization of this
vehicle must encourage both field and laboratory studies in wildlife and human
populations to satisfy the need for causal linkages.
2. This vehicle must promote innovative, multi-disciplinary research on transgenerationally-
transmitted early markers of exposure that predict long-term, delayed, loss of function.
These research efforts should be designed to determine the most sensitive endpoint(s) (the
lower-limits of effect) using a multigenerational model.
3. The proposed vehicle must promote innovative, cross-discipline, multi-level (gene to
ecosystem) research, that addresses pollution problems recognized as a result of damage
in the field from ambient levels of xenobiotic compounds.
4. This vehicle should also establish a review process for research proposals that is geared
to support the cutting edge research necessary to keep ahead of the technologies
producing new and more powerful pesticides. This must be a new review process
separate from the current practice in use today.
5. The vehicle should also fund the development of inexpensive, short-term screening
techniques to test new and old products for endocrine, nervous, and immune system
disruptive capacity.
6. This vehicle would serve to accelerate testing of banned and restricted products that still
pose a threat to humans and wildlife because of their persistence and presence in human
tissue.
7. In addition to considering human impacts directly, this vehicle should also support
exposure and effect studies using free-ranging wildlife as models for human exposure and
effects resulting from ambient levels of xenobiotic compounds.
8. Although we increasingly are beginning to understand the mechanism of action of toxic
substances on the biology of individual organisms and on sub-organismal levels of biotic
organization, the relationship of effect at these levels to effects at higher levels of biotic
organization remain obscure. The proposed research vehicle should stimulate multi-
disciplined research relating the effects of toxic substances on individual organisms to
effects on populations, communities, and ecosystems.
115
-------
Research Needs for the Mechanisms of Action of Xenohiotic Substances
1. There are a multiple of possible deleterious endpoints from xenobiotic exposure other
than cancer. Research on diverse mechanisms of effect and the multiplicity of
biological endpoints must be increased.
2. Some effects of xenobiotics are insidious, long-term, and multigenerational. An
increase in long-term studies of single exposure, low-dose, or embryonic and
developmental exposure is warranted.
3. The lower-limits of effects are unknown for virtually all chemicals, especially
considering long-term and multigenerational studies. The establishment of lower-
thresholds for all known effects must be undertaken.
4. Central to our establishment of guidelines for chemical usage and risk assessment is
the understanding of the range of thresholds and effects within genetically diverse
populations, and not merely the mean threshold levels for effects. The identification
of thresholds for "sensitive" members of populations is warranted for future risk
assessment decisions.
5. There are large gaps in our knowledge concerning the effects of xenobiotics in diverse
groups of organisms, such as reptiles, amphibians, chondrichthian fishes (sharks, skates
and rays), invertebrates and vascular plants. These groups form important parts of the
food web and habitats they live in and, and many are showing world-wide declines,
amphibians and sharks. An increased research emphasis is needed in these groups.
6. Wildlife and humans are exposed to a large diversity of chemicals. The interactions
of multiple xenobiotic chemicals must be investigated in order to elucidate possible
synergisms or antagonisms.
7. The influence of environmental factors, such as temperature, pH, salinity, and
dissolved oxygen content are poorly understood with regards to how they modify
xenobiotic toxicities. The study of environmental factors for diverse habits, such as
warm-water lakes, estuaries, and tropical marches are clearly warranted.
116
-------
2.2.10 Acknowledgements
The "Eutrophication" section and the "Eutrophication Case Study" of this report are based
on a background paper on "effects of nutrients on coastal water quality" prepared by R. W.
Howarth for the Committee on Wastewater Management for Coastal Urban Areas, Water Science
and Technology Board, National Research Council. This background paper forms the basis of
Appendix 1 of the report of the Committee, in review, and is copyrighted by the National
Academy of Sciences. Portions are used here with permission.
The authors of this document wish to express their sincere appreciation to Ms. Lisa Reyes
in acknowledgement of her exceptional efforts in assembling the various portions of this text into
its final form.
117
-------
2.2.11
Animal Species
Beluga
English sole
Rock sole
Starry flounder
Flathead sole
White croaker
White perch
Windowpane flounder
Winter flounder
Bullhead trout
Atlantic croaker
California halibut
Dolly Varden
Hornyhead turbot
Pacific halibut
Herring gull
Forster's tern
Ring-billed gull
Western gull
California gull
Pink salmon
Coho salmon
Chinook salmon
Striped bass
Sea star
Atlantic cod
Rainbow trout
Polychaete
Mussel
Caspian tern
Bald eagle
Black-crowned night-heron
Common tern
Double-crested cormorant
Osprey
Mink
Otter
Lake trout
Common snapping turtle
Great blue heron
Virginia rail
Delphinapterus leucas
Parophrys vetulus
Lepidopsetta bilineata
Platickthys stellatus
Hippoglossoides elassodon
Genyonemus lineatus
Morone americana
Scopthalmus aquosus
Pseudopleuronectes americanus
Salvelinus confluentus
Micropogonias undulatus
Paralichthys californicus
Salvelinus malma
Scophthalmus maximus
Hippoglossus sp.
Larus argentatus
Sterna forsteri
Larus delawarensus
Larus occidentalis
Larus californicus
Onchorhynchus gorbuscha
Onchorhynchus Idsutch
Oncorhynchus tshawytscha
Morone saxatilus
Asterias rubens
Gadus morhua
Salmo gairdneri
Nereis virens
Mytilus edulis L.
Hydroprogne caspia
Haliaetus leucocephalus
Nycticorax
Sterna hirundo
Phalactrocorax auritus
Pandion haliaetus
Mustela vison
Lutra canadensis
Sylvelinus namaycush
Chelydra serpentina
Ardea herodias
Ralus virginianus
118
-------
2.2.12 REFERENCES
Abe, S. Inoue, Y., and Takamatsu, M. 1975. Polychlorinated biphenyl residues in plasma of
Yusho children born to mothers who had consumed oil contaminated by PCS. Acta
Medica Fukuoka 66: 605-609.
Abel, E.L. 1984. Fetal alcohol syndrome and fetal alcohol effects. New York, NY: Plenum Press.
Aber, J., Nadelhoffer, K., Steudler, P., and Meiillo, J. 1989. Nitrogen saturation in northern forest
ecosystems. BioScience 39: 378-386.
Abbot, B.D., Harris, M.W., and Birabaum, L.S. 1992. Comparisons of the effects of TCDD and
hydrocortisone on growth factor expression provide insight into their interaction in the
embryonic mouse palate. Teratology 45(1): 35-53.
Acey, R., Healy, P., Unger, T.F., Ford, C.E., and Hudson, RA. 1987. Growth and aggregation
behavior of representative phytoplankton as affected by the environmental contaminant
Di-n-butyl Phthalate. Bulletin of Environmental Contamination and Toxicology 39:1-6.
Addison, R.F. 1989. Organochlorines and marine mammal reproduction. Canadian Journal of
Fisheries and Aquatic Science 46: 360-368.
Agency for Toxic Substances and Disease Registry (ATSDR). U.S. Public Health Service.
Toxicological Profile for Benzo[a]Pyrene. Draft. October 1987.
Agency for Toxic Substances and Disease Registry (ATSDR). 1987. Toxicological Profile for
Tetrachloroethylene. Draft. U.S. Public Health Service. Oak Ridge National Laboratory.
December.
»
Agency for Toxic Substances and Disease Registry (ATSDR). 1987. Toxicological Profile for
Aldrin/Dieldrin. Draft. U.S. Public Health Service. Oak Ridge National Laboratory.
November.
Agency for Toxic Substances and Disease Registry (ATSDR). 1987. Toxicological Profile for
2,3,7,8-Tetrachlorodibenzo-p-Dioxin. Draft. U. S. Public Health Service. Oak Ridge
National Laboratory. December.
Agency for Toxic Substances and Disease Registry (ATSDR). 1987. Toxicological Profile for
Cadmium. Draft. U.S. Public Health Service. Oak Ridge National Laboratory. November.
Agency for Toxic Substances and Disease Registry (ATSDR). 1987. Toxicological Profile for
Selected PCBs (Arochlor-1260, -1254, 1248, 1242, 1232, 1221, and -1016. (Draft).
U.S. Public Health Service. Oak Ridge National Laboratory. November.
119
-------
Agency for Toxic Substances and Disease Registry (ATSDR). 1987. lexicological Profile for
Di(2-ethylhexyl)Phthalate. Draft. U.S. Public Health Service. Oak Ridge National
Laboratory. December.
Agency for Toxic Substances and Disease Registry (ATSDR). 1988. Toxicological Profile for
Mercury. Draft. U.S. Public Health Service. Oak Ridge National Laboratory. December.
Agency for Toxic Substances and Disease Registry (ATSDR). 1988. Toxicological Profile for
Chlordane. Draft. U.S. Public Health Service. Oak Ridge National Laboratory. December.
Agency for Toxic Substances and Disease Registry (ATSDR). 1988. Toxicological Profile for
DDT, DDE, and ODD. Draft. U.S. Public Health Service. Oak Ridge National Laboratory.
December.
Agency for Toxic Substances and Disease Registry (ATSDR). 1988. Toxicological Profile for
Pentachlorophenol. Draft. U.S. Public Health Service. Oak Ridge National Laboratory.
December.
Agency for Toxic Substances and Disease Registry (ATSDR). 1988. Toxicological Profile for
Lead. Draft. U.S. Public Health Service. Oak Ridge National Laboratory. February.
Agency for Toxic Substances and Disease Registry (ATSDR). 1989. United States Department
of Public Health. Toxicological Profile for 2,3,7,8-Tetrachlorodibenzo-p-Dioxin. June.
Ahlborg, U., Hanberg, A., and Kenne, K. 1992. Risk assessment of polychlorinated biphenyls
(PCBs). Pp. 86. Institute of Environmental Medicine. Kardinska Institute!. Stockholm,
Sweden. Nord 1992: 26.
Albertini, S., Friederich, U., Holderegger, C, and Wurgler, F.E. 1988. The in vitro porcine brain
tubulin assembly assay: effects of a genotoxic carcinogen (aflatoxin Bl), eight tumor
promoters and nine miscellaneous substances. Mutation Research 201(2): 283-292.
Allen, J., Van Miller, J., and Norback, D. 1975. Tissue distribution, excretion and biological
effects of (14C) tetrachlorodibenzo-p-dioxin in rats. Food Cosmetology and Toxicology
13: 501-505.
Allen, J.R. and Barsotti, DA. 1976. The effects of transplacental and mammary movement of
PCBs on infant Rhesus monkeys. Toxicology 6(3): 331-340.
Allen, J.R., Barsotti, DA., Van Miller, J., Abrahamson, L., and Lalich, L. 1977. Morphological
changes in monkeys consuming a diet containing low levels of 2,3,7,8-
tetrachlorodibenzo-p-dioxin. Food Cosmetology and Toxicology 15(5): 401-410.
120
-------
Allen, J.R., Barsotti, DA., Lambrecht, L., and Van Miller, J. 1979. Reproductive effects of
halogenated aromatic hydrocarbons on nonhuman primates. Annals of the New York
Academy of Sciences 320: 419-425.
Amdur, M.O., Doull, J., and Klaassen, CD. 1991. Casarett and Doull's toxicology: the basic
science of poisons. New York, NY: Pergamon Press.
Anders, M.W. 1985. Bioactivation of Foreign Compounds. New York, NY: Academic Press.
Andersen, T., Schartau, A., and Paasche, E. 1991. Quantifying external and internal nitrogen and
phosphorus pools, as well as nitrogen and phosphorus supplied through remineralization,
in coastal marine plankton by means of a dilution technique. Marine Ecology Progress
Series 69: 67-80.
Anderson, D. 1989. Toxic algal blooms and red tides: a global perspective, in: Okaichi, T.,
Anderson, D.M., and Nemoto, T. (eds.). Red Tides: Biology, Environmental Science and
Toxicology. Elsevier, New York.
Andersson, L., Nikolaidis, E., Brunstrom, B., Bergman, A., and Dencker, L. 1991. Effect of
polychlorinated biphenyls with Ah receptor affinity on lymphoid development in the
thymus and the bursa of Fabricius of chick embryos in ovo and in mouse thymus anlagen
m vitro. Toxicology and Applied Pharmacology 107: 183-188.
Ando, M., Hirano, S., and Itoh, Y. 1985. Transfer of hexachlorobenzene (HCB) from mother to
new-born baby through placenta and milk. Archives of Toxicology 56(3): 195-200.
Andren, A., and Strand, J. 1981. Atmospheric deposition of particulate organic carbon and PAHs
to Lake Michigan. Pp. 459-479. in: SJ. Eisenreich (ed.). Atmospheric Inputs of Pollutants
to Natural Waters. Ann Arbor Press.
Andres, P. 1984. IgA-IgG disease in the intestine of Brown-Norway rats ingesting mercuric
chloride. Clinical Immunology and Immunopathology 30: 488-494.
Annau, Z. and Cuomo, V. 1988. Mechanisms of neurotoxicity and their relationship to behavioral
changes. Toxicology 49: 219-229.
Anner, B.M. and Moosmayer, M. 1992. Mercury inhibits Na-K-ATPase primarily at the
cytoplasmic side. American Journal of Physiology 262(5 pt 2): F843-848.
Antunes-Madeira, M.C. and Madeira, V.M. 1990. Membrane fluidity as affected by the
organochlorine insecticide DDT. Biochimica et Biophysica Acta 1023(3): 469-474.
121
-------
Aoyama, T., Gelboin, H.V., and Gonzalez, FJ. 1990. Mutagenic activation of 2-amino-3-
methylimidazo[4,5-f]quinoline by complementary DNA-cxpresscd human liver P-450.
Cancer Research 50: 2060-2063.
Aschner, M. and Aschner, H.L. 1990. Mercury neurotoxicity: mechanisms of blood-brain barrier
transport. Neuroscience and Biobehavioral Review 14: 169-176.
Astroff, B., and Safe, S. 1990. 2,3,7,8-tetrachlorodibenzo-p-dioxin as an antiestrogen: effect on
rat uterine peroxidase activity. Biochemical Pharmacology 39: 485-488.
Astroff, F., Rowlands, C, Dickerson, R., and Safe, S. 1990.2,3,7,8-tetrachorodibcnzo-p-dioxin
inhibition of 17 beta-estradiol-induced increases in rat uterine epidermal growth factor
receptor binding activity and gene expression. Molecular and Cellular Endocrinology
72(3): 247-252.
Astroff, F., Eldridge, B., and Safe, S. 1991. Inhibition of the 17 beta-estradiol-induced and
constitutive expression of the cellular protooncogene c-fos by 2,3,7,8-tetrachorodibenzo-
p-dioxin (TCDD) in the female rat uterus. Toxicology Letters 56(3): 305-315.
Astroff, F. and Safe, S. 1991. 6-Alkyl-l,3,8-trichlorodibenzofurans as antiestrogens in female
Sprague-Dawley rats. Toxicology 69: 187-97.
Atchison, W.D. and Narahashi, T. 1982. Methylmercury-induced depression of neuromuscular
transmission in the rat. Neurotoxicology 3: 37-50.
Atchison, W.D. 1986. Extracellular calcium-dependent and independent effects of methylmercury
on spontaneous and potassium-evoked release of acetylcholine at the neuromuscular
junction. Journal of Pharmacology and Experimental Therapy 237: 672-680.
Austin, A. and Munteanu, N. 1984. Evaluation of changes in a large oligotrophic wilderness park
lake exposed to mine tailing effluent for 14 years: the phytoplankton. Environmental
Pollution (Series A) 33: 39-62.
Austin, A.P., Harris, G.E., and Lucey, W.P. 1991. Impacts of an organophosphate herbicide
(Glyphosate) on periphyton communities developed in experimental streams. Bulletin of
Environmental Contamination and Toxicology 47: 29-35.
Baden, S.P., Loo, L.-O., Pihl, L., and Rosenberg, R. 1990. Effects of eutrophication on benthic
communities including fish: Swedish west coast. Ambio 19: 113-122.
Baker, J.R. 1989. Pollution-associated uterine lesions in grey seals from the Liverpool Bay area
of the Irish Sea. Veterinary Record 125: 303.
122
-------
Ball, L.M., and Chabra, R.S. 1981. Intestinal absorption of nutrients in rats treated with 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD). Journal of Toxicology and Environmental Health
8: 629-638.
Ballschmiter, K., Zell, M., and Neu, H J. 1978. Persistence of PCBs in the ecosphere: will some
PCS components 'never1 degrade? Chemosphere 2: 173-176.
Ballschmiter, K., Buchert, H., Bikler, S., and Zell, M. 1981. Baseline studies of the global
pollution: IV. The pattern of pollution by organochlorine compounds in the North Atlantic
as accumulated by fish. Fresenius Zeitung Analitische Chemie 306: 323-339.
Banerjee, B., Ramachandran, M., and Hussain, Q. 1986. Sub-chronic effect of DDT on humoral
immune response in mice. Bulletin of Environmental Contamination and Toxicology 37:
433-440.
Banerjee, B. 1987a. Effects of sub-chronic DDT exposure on humoral and cell-mediated
immune responses in albino rats. Bulletin of Environmental Contamination and
Toxicology 39: 827-834.
Banerjee, B. 1987b. Sub-chronic effect of DDT on humoral immune response to a thymus-
independent antigen (bacterial lipo-polysaccharide) in mice. Bulletin of Environmental
Contamination and Toxicology 39: 822-826.
Bannister, R. and Safe, S. 1987. Synergistic interactions of 2,3,7,8-TCDD and 2,2',4,4',5,5'-
hexachlorobiphenyl in C57BL/6J and DBA/2J mice: role of the Ah receptor. Toxicology
44: 159-169.
Banse, K. 1990. Does iron really limit phytoplankton production in the offshore subarctic
Pacific? Limnology & Oceanography 35: 772-775.
Barnett, J., Holcomb, D., Menna, J., and Soderberg, L. 1985. The effect of prenatal chlordane
exposure on specific anti-influenza cell-mediated immunity. Toxicology Letters 25(3):
229-238.
Barnett, J., Barfield, L., Walls, R., Joyner, R., Owens, R., and Soderberg, L. 1987. The effect of
in utero exposure to hexachlorobenzene on the developing immune response of BALB/c
mice. Toxicology Letters 39(2-3): 263-274.
Barsotti, D.A., Abrahamson, L.J., and Allen, J.R. 1979. Hormonal alterations in female Rhesus
monkeys fed a diet containing 2,3,7,8-tetrachlorodibenzo-p-dioxin. Bulletin of
Environmental Contamination and Toxicology 21: 463-469.
Bartdlome, J., Trepanier, P., Chait, E.A., Seilder, F.J., Dcskin, R., and Slotkin, TA. 1982.
Neonatal methylmercury poisoning in the rat: effects on development of jcentral
123
-------
catecholamine neurotransmitter systems. Toxicology and Applied Pharmacology 65: 92-
99.
Batty, J., Leavitt, RA., Biondo, N., and Polin, D. 1990. An ecotoxicological study of a
population of the white footed mouse (Peromyscus leucopus) inhabiting a polychlorinated
biphenyls-contaminated area. Archives of Environmental Contamination and Toxicology
19: 283-290.
Baukloh, V., Bohnet, H., Trapp, M., Heeschen, W., Feichtinger, W., and Kemeter, P. 1985.
Biocides in human follicular fluid. Annals of the New York Academy of Sciences 442:
240-250.
Baumann, P. and Harshberger, J. 1985. Frequencies of liver neoplasia in a feral population and
associated carcinogens. Marine Environmental Research 17: 324-327.
Baumann, P.C., Smith, W.D., and Parkland, W.K. 1987. Tumor frequencies and contaminant
concentrations in brown bullheads from an industrialized river and a recreational lake.
Transactions of the American Fisheries Society 116: 79-86.
Beeton, A.M. 1965. Eutrophication of the St. Lawrence Great Lakes. Limnology &
Oceanography: 240-254.
Beland, P., Vezina, A., and Martineau, D. 1988. Potential for growth of the St. Lawrence
(Quebec, Canada) beluga whale (Delphinaptems leucas) population based on modelling.
Journal du Conseil. Conseil International Pour UExploration De La Mer. 45: 22-32.
Beland, P., DeGuise, S., Girard, C, Lagase, A., Martineau, D., Michaud, R., Muir, D., Norstrom,
R., Pelletier, E., and Shugart, L. 1991. Toxic compounds and health and reproductive
effects in St. Lawrence beluga whales. Pp. 26-27 in: Schneider, S. and Campbell, R.
(eds.). Cause-Effects Linkages II Symposium Abstracts. Michigan Audubon Society,
Lansing, MI.
Beland, P., DeGuise, S., and Plante, R. 1992. Toxicology and pathology of St. Lawrence marine
mammals. Report SLNffi, 3872 Parc-Lafontaine, Montreal, H2L 3M6. Wildlife
Toxicology Fund, World Wildlife Fund Canada.
Bellon, B., Capron, M., Druet, E., Verroust, P., Vial, M.C., Sapin, C., Girard, J.F., Foidart, J.M.,
Mahieu, P., and Druet, P. 1982. Mercuric chloride induced autoimmune disease in
Brown-Norway rats: Sequential search for anti-basement membrane antibodies and
circulating immune complexes. European Journal of Clinical Investigation 12: 127-133.
Beggs, M., Menna, J., and Barnett, J. 1985. Effect of chlordane on influenza type A virus and
herpes simplex type 1 virus replication in vitro. Journal of Toxicology and Environmental
Health 16(2): 173-188.
124
-------
Bellinger, D., Leviton, A., Wateraaux, C, Needleman, H., and Rabinowitz, M. 1987.
Longitudinal analyses of prenatal and postnatal lead exposure and early cognitive
development. New England Journal of Medicine 316: 1037-1043.
Benvenue, A., Ogata, J.N., and Hylin, J.W. 1972. Organochlorine pesticides in rainwater. Oahu,
Hawaii. 1971-1972. Bulletin of Environmental Contamination and Toxicology 8: 238-
241.
van den Berg, M., Heeremans, C., Veenhoven, E., and Olie, K. 1987. Transfer of
polychlorinated dibenzo-p-dioxins and dibenzofurans to fetal and neonatal rats.
Fundamental and Applied Toxicology 9: 635-644.
van den Berg, K., Zurcher, C., and Brouwer, A. 1988a. Effects of 3,4,3',4'-tetrachlorobiphenyl
on thyroid function and histology in marmoset monkeys. Toxicology Letters 41: 77-86.
van den Berg, K., Brouwer, A., and van Bekkum, D. 1988b. Chronic toxicity of 3,4,3',4'-
tetrachlorobiphenyl in the marmoset monkey (Callithrix jacchus). Toxicology 48: 209-
224.
Berger, D. and Mueller, J. No date. Ospreys in Northern Wisconsin. N.P. 340-341.
Bergman, A., and Olsson, M. 1985. Pathology of baltic grey seal and ringed seal females with
special reference to adrenocortical hyperplasia: is environmental pollution the cause of
widely distributed disease syndrome? Finnish Game Research 44: 47-62.
Bernaudin, J.f., Druet, E., Druet, P., and Masse, R. 1981. Inhalation or ingestion of organic or
inorganic mercurial produes auto-immune disease in rats. Clinical Immunology and
Immunopathology 20: 129-135.
Berry, J.W., D. W. Osgood, and St. John, P.A. 1974. Chemical villains: a biology of pollution.
St. Louis, MO: C.V. Mosby Co.
Bestervelt, L.L., Nolan, C.J., Cai, Y., Maimansomsuk. P., Mousigian. CA., and Piper, W.N.
1991. Tetrachlorodibenzo-p-dioxin alters rat hypothalamic endorphin and \JL opioid
receptors. Neurotoxicology and Teratology 13(5): 495-497.
Bidelman, J.F. and Olney, C.E. 1974. Chlorinated hydrocarbons in the Sargasso Sea atmosphere
and surface water. Science 183: 516-518.
Biegel, L. and Safe, S. 1990. Effects of 2,3,7,8-tetrachorodibenzo-p-dioxin (TCDD) on cell
growth and the secretion of estrogen-induced 34-, 52-, and 160-kDa proteins in human
breast cancer cells. Journal of Steroid Biochemistry and Molecular Biology 37(5): 723-
732.
125
-------
Biessmann, A. 1982. Effects of PCBs on gonads, sex hormone balance and reproduction
processes of Japanese quail (Coturnix coturnix japonica) after ingestion during sexual
maturation. Environmental Pollution (Series A). 27: 15-30.
Biggs, D.C., Rowland, R.G., O'Connors, H.B., Jr., Powers, CD., and Wurster, C.F. 1978. A
comparison of the effects of chlordane and PCB on the growth, photosynthesis, and cell
size of estuarine phytoplankton. Environmental Pollution 14: 253-263.
Binder, R.L. and Lech, J J. 1984. Xenobiotics in gametes of Lake Michigan lake trout (Salvelinus
namaycush) induce hepatic monoxygenase activity in their offspring. Fundamental
Applied Toxicology 4: 1042-1054.
Birobaum, L.S., Weber, H., Harris, M., Lamb, J., and McKinney, J. 1985. Toxic interaction of
specific polychlorinated biphenyls and 2,3,7,8-tetrachlorodibenzo-p-dioxin: increased
incidence of cleft palate in mice. Toxicology and Applied Pharmacology 77: 292-302.
Bishop, CA., Brooks, RJ., Carey, J.H., Ng, P., Norstrom, RJ., and Lean, D.R.S. 1991. The case
for a cause-effect linkage between environmental contamination and development in eggs
of the common snapping turtle (Chelydra S. Serpentina) from Ontario, Canada. Journal
of Toxicology and Environmental Health 33(4): 521-548.
Bitman, J. and Cecil, H.C. 1970. Estrogenic activity of DDT analogs and polychlorinated
biphenyls. Journal of Agricultural and Food Chemistry 18: 1108-1112.
Black, J., Dymerski, P., and Zapisek, W. 1981. Environmental carcinogenesis studies in the
western New York Great Lakes aquatic environment. Pp. 215-225 in: Branson and
Dickson (eds.). Aquatic Toxicology and Hazard Assessment. Fourth Conference, ASTM
STP 377. American Society for Testing and Materials.
Black, J., Evans, E., Harshberger, J., and Ziegel, R. 1982. Epizootic neoplasms in fishes from a
lake polluted by copper mining wastes. Journal of the National Cancer Institute 69(4):
915-926.
Black, J.J., Fox, H., Black, P., and Block, F. 1985. Carcinogenic effects of river sediment
extracts in fish and mice. Pp. 415-427 in: Jolley, R.L., Bull, RJ., Davis, W.P., Katz, S.,
Roberts, Jr., M.H., Jacobs, VA. (eds.). Water chlorination: chemistry, environmental
impact and health effects. Chelsea, Michigan: Lewis Publishers, Inc. (Government
Canada).
Blakely, B,R. 1988. Humoral immunity in aged mice exposed to cadmium. Canadian Journal of
Veterinary Research 52: 291-292.
126
-------
Blaylock, B.L., Soderberg, L.S.F., Gandy, J., Menna, J.H., Denton, R., and Barnett, J.B. 1990.
Cytotoxic T-lymphocyte and NK responses in mice treated prcnatally with chlordane.
Toxicity Letters 51: 41-49.
Blaylock, B.L., Holladay, S.D., Comment, CE., Heindel, JJ., and Luster, M.I. 1992. Exposure
to tetrachlorodibenzo-p-dioxin (TCDD) alters fetal thymocyte maturation. Toxicology and
Applied Pharmacology 112(2): 207-213.
Biokpoel, H., and Tessier, G. 1986. The ring-billed gull in Ontario: a review of a new problem
species. Canadian Wildlife Service, Occasional Paper No. 57. CW69-1/57-1986 E.
Biokpoel, H. 1988. Status of colonial waterbirds nesting on Lake Ontario in 1987. Presented at
the 31st Conference on Great Lakes Research, Hamilton, Ontario. May.
Boadi, W.Y., Urbach, J., Barnea, E.R., Brandes, J.M., and Yannai, S. 1991. In vitro effect of
mercury on aryl hydrocharbon hydroxylase, quinone reductase, catecholamine-O-
methyltransferase and glucose-6-phosphate. Pharmacology and Toxicology 68:317-321.
Boadi, W.Y., Urbach, J., Brandes, J.M., and Yannai, S. 1992. In vitro effect of mercury on
enzyme activities and its accumulation in the first-trimester human placenta.
Environmental Research 57: 96-106.
den Boer, M. 1983. Reproduction decline of harbour seals: PCBs in the food and then- effect on
mink. Pp. 77-86 in: van Rossum, T. (e&). Annual Report RIN. Leersum, The Netherlands:
Rijksinstituut voor Naturrbeheer.
Boersma, D.C., Ellenton, J.A., and Yagminas, A. 1986. Investigation of the hepatic mixed-
function oxidase system in herring gull embryos in relation to environmental pollutants.
Environmental Toxicology and Chemistry 5: 309-318.
Bondi, S.C. and McKee, M. 1991. Disruption of the potential across the synaptosomal plasma
membrane and mitochndria by neurotoxic agents. Toxicology Letters 58: 13-21.
Bonithon-Kopp, C, Huel, G., Grasmick, C, Sannini, H., Moreau, T., and Wendling, R. 1986a.
Prenatal exposure to lead and cadmium and psychomotor development of the child at 6
years. Neurobehavioral Toxicology and Teratology 8(3): 307-310.
Bonithon-Kopp, C., Huel, G., Grasmick, C., Sannini, H., and Moreau, T. 1986b. Effects of
pregnancy on the inter-individual variations hi blood vessels of lead, cadmium, and
mercury. Biological Research in Pregnancy 7(1): 37-42.
Bookstaff, R.C., Kamel, F., Moore, R.W., Bjerke, D.L., and Peterson, R.E. 1990. Altered
regulation of pituitary gonadotropin-releasing hormone (GnRH) receptor number and
127
-------
pituitary responsiveness to GnRH in 2,3,7,8-tetrachlorodibenzo-p-dioxin-treated male
rats. Toxicology and Applied Pharmacology 105(1): 78-92.
Bookstaff, R.C., Moore, R.W., and Peterson, RE. 1990. 2,3,7,8-tetrachorodibenzo-p-dioxin
increases the potency of androgens and estrogens as feedback inhibitors of luteinizing
hormone secretion in male rats. Toxicology and Applied Pharmacology 104: 212-224.
Borlakoglu, J.T., Edwards-Webb, DJ., and Dils, R.R. 1990. Polychlorinated biphenyls increase
fatty acid desaturation in the proliferating endoplasmic reticulum of pigeon and rat livers.
European Journal of Biochemistry 188(2): 327-332.
Borlakoglu, J.T., Stegeman, J., and Dils, R.R. 1991. Induction of hepatic cytochrome P-4501A1
in pigeons treated in vivo with Aroclor 1254, a commercial mixture of polychlorinated
biphenyls (PCBs). Comparative Biochemistry and Physiology 99(3): 279-288.
Borowitzka, M.A. 1972. Intertidal algal species diversity and the effect of pollution. Australian
Journal of Marine and Freshwater Science 23: 73-84.
Borrell, A., and Aguilar, A. 1991. Pollution by PCBs in striped dolphins affected by the western
Mediterranean epizootic. Pp. 121-127 in: Pastor, X. and Simmonds, M. (eds.). The
Mediterranean Striped Dolphin Die-Off. Proceedings of the Mediterranean striped
dolphin mortality International Workshop, Palma de Mallorca, 4-5 November, 1991.
Bomhausen, M., Musch, H.R., and Greim, H. 1980. Operant behaviour performance changes in
rats after prenatal methyl-mercury exposure. Toxicology and Applied Pharmacology 56:
305-310.
Bowerman, W.B., Best, D., Kubiak, T., Postupalsky, S., and Tillitt, D. 1991. Bald eagle
reproduction impairment around the Great Lakes: association with organochlorine
contamination. Pp. 31-32 in: Schneider, S. and Campbell, R. (eds.). Cause-Effect
Linkages II Symposium Abstracts. Michigan Audubon Society, September 27-28, 1991.
Bowes, G.W., and Jonkel, CJ. 1975. Presence and distribution of polychlorinated biphenyls
(PCB) in arctic and subarctic food chains. Journal of Fisheries Board of Canada 32:
2111-2123.
Bowman, R.E., Schantz, S.L., Gross, M.L., and Ferguson, S.A. 1989. Behavioral effects in
monkeys exposed to 2,3,7,8-TCDD transmitted maternally during gestation and for four
months of nursing. Chemosphere 18: 235-242.
Boynton, W.R., Kemp, W!M., and Osborne, C.G. 1980. Nutrient fluxes across the sediment-
water interafae in the turbid zone of a coastal plain estuary. In: V.S. Kennedy, (ed.).
Estuarine Perspectives. New York, NY: Academic Press.
128
-------
Boynton, W.R., Kemp, W.M., and Keefe, C.W. 1982. A Comparative analysis of nutrients and
other factors influencing estuarine phytoplankton production. In: Kennedy, V.S. (ed).
Estuarine Perspectives. New York, NY: Academic Press.
Bozelka, B., and Salvaggio, J. 1985. Immunomodulation by environmental contaminants:
asbestos, cadmium, and halogenated biphenyls: a review. Environmental Carcinogenesis
Reviews 3(1): 1-62.
Bradlaw, J. and J. Casterline, Jr. 1979. Induction of enzyme activity in cell culture: a rapid
screen for detection of planar polychlorinated organic compounds. Journal of Associations
of Official Analytical Chemistry 62(4): 904-926.
Brewster, D.W. and Matsumura, F. 1988. Reduction of adipose tissue lipoprotein lipase activity
as a result of in vivo administration of 2,3,7,8-tetrachlorodibenzo-p-dioxin to the guinea
pig. Biochemical Pharmacology 37(11): 2247-2253.
Bricelj, M, and Kuenstner, S. 1989. The feeding physiology and growth of bay scallops and
Mussels. In: Cosper, E.M., Carpenter, E.J., and Bricelj, V.M. (eds.). Novel phytoplankton
blooms: causes and impacts of recurrent nrown tides and other unusual blooms. Lecture
Notes on Coastal and Estuarine Studies. Berlin: Springer-Verlag.
Britt, J.O., and Howard, E.B. 1983. Tissue residues of selected environmental contaminants in
marine mammals. Pp. 80-94 in: Howard, E.B. (ed.). Pathobiology of Marine Mammal
Diseases. Boca Raton, FL: CRC Press.
Brooks, R. 1987. Snapping turtles (Chelydra serpentind) as biomonitors of organochlorine
contamination in wetlands. Toxicology Fund Progress Report. July 10, 1987.
Brouwer, A., Reijnders, P.J.H., and Koeman, J.H. 1989. Polychlorinated biphenyl (PCB>-
contaminated fish induces vitamin A and thyroid deficiency in the common seal, (Phoca
vitulina). Aquatic Toxicology 15: 99-106.
Bryson, P.D. 1989. Comprehensive Review in Toxicology. Rockville, Maryland. Aspen
Publishers.
Buchmuller-Rouiller, Y., Ransijn, A., and Mauel, J. 1989. Lead inhibits oxidative metabolism
of macrophages exposed to macrophage-activating factor. Biochemistry Journal 260:325-
332.
Buff, K. and Brundl, A. 1992. Specific binding of polychlorinated biphenyls to rat liver cytosol
protein. Biochemical Pharmacology 43(5): 965-970.
Bulger, W.H., Muccitelli, R.M., and Kupfer, D. 1978a. Interactions of chlorinated hydrocarbon
pesticides with the 8S estrogen-binding protein in rat testes. Steroids 32: 165-177.
129
-------
Bulger, W.H., Muccitelli, R.M., and Kupher, D. 1978b. Studies on the induction of rat uterine
ornithine decarboxylase by DDT analogs, n. Kinetic characteristics of ornithine
decarboxylasc induced by DDT analogs and estradiol. Pesticide Biochemistry and
Physiology 8: 263-270.
Bulger, W.H., Muccitelli, R.M., and Kupher, D. 1978c. Studies on the MI vivo and in vitro
estrogenic activities of methoxychlor and its metabolites: role of hepatic mono-oxygenase
in methoxychlor activation. Biochemical Pharmacology 27: 2417-2423.
Bulger, W.H. and Kupfer, D. 1978. Studies on the induction of rat uterine omithine
decarboxylase by DDT analogs. I. Comparison with estradiol-17B activity. Pesticide
Biochemistry and Physiology 8: 165-177.
Bulger, W. and Kupfer, D. 1983. Effect of xenobiotic estrogens and structurally related
compounds on 2-hydroxylation of estradiol and on other monoxygenase activities in rat
liver. Biochemical Pharmacology 32(6): 1005-1010.
Bulger, W.H. and Kupfer, D. 1983a. Estrogenic action of DDT analogs. American Journal of
Industrial Medicine 4: 163-173.
Bulger, W.H. and Kupfer, D. 1983b. Effect of xenobiotic estrogens and structurally related
compounds on 2-hydroxylation of estradiol and on other monooxygenase activities in rat
liver. Biochemical Pharmacology 32(6): 1005-^1010.
Burbacker, T.M., Monnett, C, Grant, K.S., and Mottet, N.K. 1984. Methylmercury exposure and
reproductive dysfunction in the nonhuman primate. Toxicology and Applied
Pharmacology 75: 18-24.
Burger, J. 1990. Behavioral effects of early postnatal lead exposure in herring gull (Larus
argentatus) chicks. Pharmacology, Biochemistry, and Behavior 35: 7-13.
Burkholder, J.M., Noga, E.J., Hobbs, C.H., and Glasgow, H.B. 1992a. New "phantom"
dinoflagellate is the causative agent of major estuarine fish kills. Nature 358: 407-410.
Burkholder, J.M., Mason, KM., and Glasgow, H.B. 1992b. Water-column nitrate enrichment
promotes decline of eelgrass Zostera marina: evidence from seasonal mesocosm
experiments. Marine Ecology Progress Series 81: 163-178.
Bush, B., Snow, J., and Koblintz, R. 1984. Polychlorobiphenyl (PCB) congeners, p,p'-DDE, and
hexachlorobenzene, and hexchlorobenzene in maternal and fetal cord blood from mothers
in upstate New York. Archives of Environmental Contamination and Toxicology 13:517-
527.
130
-------
Bush, B., Snow, J., Connor, S., and Koblintz, R. 1985. Polychlorinated biphenyl congeners
(PCBs), p,p'-DDE and hexachlorobenzene in human milk in three areas of upstate New
York. Archives of Environmental Contamination and Toxicology 14: 443-450.
Bush, B., Bennett, A., and Snow, J. 1986. Polychlorobiphenyl congeners, p,p'-DDE, and sperm
function in humans. Archives of Environmental Contamination and Toxicology 15: 333-
341.
Bush, B., Bennett, A., and Snow, J. 1990. Pharmacodynamics of PCB congeners in the brain of
the rat and monkey. Paper No. 407 presented at the SETAC Annual Meeting, Global
environmental issues: challenges for the 90's, Arlington, VA.
Butler, M.A., Iwasaki, M., Guengerich, P.P., and Kadlubar, F.F. 1989. Human cytochrome P-
450pA (P-450IA2), the phenacetin O-deethylase, is primarily responsible for the hepatic
3-demethylation of caffeine and N-oxidation of carcinogenic arylamines. Proceedings of
the National Academy of Sciences U.SA. 86: 7696-7700.
Butt, A.J. 1992. Numerical models and nutrient reduction strategies in Virginia. Coastal
Management. 20: 25-36.
Byrne, J.J., Carbone, J.P., and Hanson, E.A. 1987. Hypothyroidism and abnormalities in the
kinetics of thyroid hormone metabolism in rats treated chronically with polychlorinated
biphenyl and polybrominated biphenyl. Endocrinology 121(2): 520-527.
Cabral, J. 1985. DDT: laboratory evidence. In: Interpretation of Negative Epidemiological
Evidence for Carcinogenicity. Proceedings of Symposium, Oxford, 4-6 July 1983. Wald
and Doll (eds.). IARC Scientific Publications No. 65.
Cairns, V., and Fitzsimmons, J. 1987. The occurrence of epidermal papiilomas and liver
neoplasia in white suckers (Catostomus commersoni) from Lake Ontario. Abstract and
Presentation of Fourteenth Annual Aquatic Toxicity Workshop. November 1-4, 1987.
Calabrese, E.J. and Sorenson, A.J. 1977. The health effects of PCBs with particular emphasis on
human high risk groups. Reviews of Environmental Health 2: 285-304.
Cambridge, M.L. and McComb, A.J. 1984. The loss of seagrasses in Cockbum Sound, Western
Australia. I. The tune course and magnitude of seagrass decline in relation to industrial
development. Aquatic Botany 20: 229-242.
Cambridge, M.C., Chaffmgs, A.W., Brittan, C, Moore, L., and McComb, A.J. 1986. The loss of
seagrass in Cockburn Sound, Western Australia. II. Possible causes of seagrass decline.
Aquatic Botany 24: 269-285.
131
-------
Capelli, R., Mingatti, V., Semino, G., and Bertarini, W. 1986. The presence of mercury (total and
organic) and selenium in human placentae. The Science of the Total Environment 48(1-
2): 69-79.
Caraco, N J., Cole, J., and Likens, G.E. 1989. Evidence for sulfate-controllcd phosphorus release
from sediments of aquatic systems. Nature 341: 316-318.
Carlsen, E., Giwercman, A., Keiding, N., and Skakkebaek, N.E. 1992. Evidence for decreasing
quality of semen during past 50 years. BMJ 305: 609-613.
Carpenter, E.J., Chang, J., Cottrell, M., Schubauer, J., Paerl, H.W., Bebout, B.M., and Capone,
D.G. 1990. Re-evaluation of nitrogenase oxygen-protective mechanisms in the planktonic
marine cyanobacterium Trichodesmium. Marine Ecology Progress Series 65: 151-158.
Carpenter, S.R., Kitchell, J.F., and Hodgson, J.R. 1985. Cascading trophic interactions and lake
productivity. BioScience 35: 634-639.
Carpenter, S.R., Kitchell, J.R., and Hodgson, J.R. 1987. Regulation of lake primary production
by food web structure. Ecology 68: 1863-1876.
Carrier, F., Owens, R.A., Nebert, D.W., and Puga, A. 1992. Dioxin-dependent activation of
murine Cypla-1 gene transcription requires protein kinase C-dependent phosphorylation.
Molecular and Cellular Biology 12(4): 1856-1863.
Cautreels, W. and Van Cauwenberghe, K. 1978. Experiments on the distribution of organic
pollutants between airborne paniculate matter and the corresponding gas phase.
Atmospheric Environment 12: 1133-1141.
Cavagnelo, R.Z. 1979. The immunology of marine mammals. Developmental Comparative
Immunology 3: 245-257.
Cederwall, H. and Elmgren, R. 1990. Biological effects of eutrophication in the Baltic Sea,
particularly the coastal zone. Ambio 19: 109-112.
Chadwick, R.W., Cooper, R.L., Chang, J., Rehnberg, G.L., and McElroy, W.K. 1988. Possible
antiestrogenic activity of lindane in female rats. Journal of Biochemical Toxicology 3:
147-158.
Chang, L.W. 1977. Neurotoxic effects of mercury - a review. Environmental Research 14:
329-373.
Chasnoff, I.J., Burns, W.J., Schnoll, S.H., and Bums, JLA. 1985. Cocaine use in pregnancy. New
England Journal of Medicine 313: 666-669.
132
-------
Chen, T.T., Reid, P.C., van Beneden, R., and Sonstegard, RA. 1986. Effect of Arochlor 1254
and mirex on estradiol-induced vitellogin production in juvenile rainbow trout (Salmo
gairdneri). Canada Journal of Fisheries and Aquatic Science 43: 169-173.
Chetty, C.S., McBride, V., Sands, S., and Rajanna, B. 1990. Effects in vitro of mercury on rat
brain Mg(**)ATPase. Archives Internationales de Physiologic et de Biochimie 98: 261-
267.
Cheung, M.K. and Verity, M.A. 1985. Experimental methylmercury neurotoxicity: locus of
mercurial inhibition of brain protein synthesis in vivo and in vitro. Journal of
Neurochemistry 44: 1799-1808.
Choi, B.H., Lapham. L.W., Amin-Zaki, L,, and Al-Saleem, T. 1978. Abnormal neuronal
migration, deranged cerebral cortical organization and diffuse white matter astrocytosis
of human fetal brain. Journal of Neuropathology and Experimental Neurology 37: 719-
733.
Choi, B.H. and Lapham, L.W. 1980. Effect of meso-2-3, dimercaptosuccinic acid on
methylmercury injured human fetal astrocytes in vitro: a lapse cinematographic phase and
electron microscopic study. Federation Proceedings 39: 396.
Chowdhury, A., Venkatakrishna-Bhatt, A., and Gautam, A. 1987. Testicular changes of rats
under lindane treatment. Bulletin of Environmental Contamination and Toxicology 38(1):
154-156.
Chu, L, Villeneuve, D., Sun, C, Secours, V,. Procter, B., Arnold, E., Clegg, D., Reynolds, L.,
and Valli, V. 1986. Toxicity of toxaphene in the rat and beagle dog. Fundamental and
Applied Toxicology 7: 406-418.
Cifone, M.G., Alesse, E., Procopio, A., Paolini, R., Morrone, S., Di Eugenio, R., Santoni, G., and
Santoni, A. 1989. Effects of cadmium on lymphocyte activation. Biochemica et
Biophysica Acta 1011: 25-32.
Clark, G.C., Blank, J.A. Germolec, D.R., and Luster, MI. 1991. 2,3,7,8-Tetrachlorodibenzo-p-
dioxin stimulation of tyrosine phosphorylation in B lymphocytes: potential role in
immunosuppression. Molecular Pharmacology 39(4): 495-501.
Clausen, J., Braestrup, L., and Berg, O. 1974. The content of polychlorinated hydrocarbons in
arctic mammals. Bulletin of Environmental Contamination and Toxicology 12: 529-534.
Clement Associates. 1989a. Toxicological Profile for alpha-, beta-, gamma, and delta-
hexachlorocyclohexane. Prepared for Agency for Toxic Substances and Disease Registry,
U.S. Public Health Service, Contract 205-88-0608.
133
-------
Clement Associates. 1989b. lexicological Profile for Chlordane. Prepared for Agency for Toxic
Substances and Disease Registry, U.S. Public Health Service, Contract 205-88-0608.
Clement Associates. 1990. Toxicological Profile for Toxaphene. Prepared for Agency for Toxic
Substances and Disease Registry, U.S. Public Health Service, Contract 205-88-0608.
Cohen, J.M., and Pmkerton, C. 1966. Widespread translocation of pesticides by air transport and
rain-out, in: Gould, R.F. (ed.). Organic Pesticides in the Environment. American
Chemical Society Advances hi Chemistry Series 60: 163-176.
Cohen, S., O'Malley, B.W., and Stastny, M. 1970. Estrogenic induction of omithine
decarboxylase in vivo and in vitro. Science 170: 336-338.
Colborn, T. 1988. Great Lakes Toxics Working Paper. Submitted to the Department of the
Environment. Government of Canada.
Colborn, T. 1989. The impact of Great Lakes toxic chemicals on human health: a working paper.
Contract Report KE 144-7-6336. Health Protection branch, Department of National
Health Welfare. Ottawa, Canada.
Colborn, T. 1991. Epidemiology of Great Lakes bald eagles. Journal of Environmental Health
and Toxicology 4: 395-453.
Coles, C.D., Smith, I.E., Femhoff, P.M., and Falek, A. 1985. Neonatal neurobehavioral
characteristics as correlates of maternal alcohol use during gestation. Alcoholism 9: 454-
459.
Concas, A., Corda, M.G., Salis, M., Mulas, M.L., Milia, A., Corongiu, F.P., and Biggio, G. 1983.
Biochemical changes in the rat cerebellar cortex elicited by chronic treatment with
methylmercury. Toxicology Letters 18: 27-33.
Cone, M., Baldauf, M., Opresko, D., and Uziel, M. 1983. Chemicals identified in human breast
milk, a literature search. U.S. Environmental Protection Agency. EPA 560/5-83-009.
Connors, P., Anderlini, V., Risebrough, R., Gilbertson, M., and Hays, H. 1975. Investigations of
heavy metals in common tern populations. Canadian Field-Naturalist 89: 157-162.
Conover, M. 1984. Frequency, spatial distribution and nest attendants of supernormal clutches
in ring-billed and California gulls. The Condor 86: 467-471.
Conover, M.R. and Hunt, G.L. 1984a. Female-female pairing and sex ratios in gulls: an
historical perspective. Wilson Bulletin 96(4): 619-625.
134
-------
Conover, MR. and Hunt, G.L. 1984b. Experimental evidence that female-female pairs in gulls
result from a shortage of breeding males. The Condor 86: 467-471.
Cosper, E.M., Dennison, W.C., Caipenter, E.J., Bricelj, V.M., Mitchell, J.G., Kuenstner, S.,
Colflesh, D.C., and Dewey, M. 1987. Recurrent and persistent "Brown Tide" blooms
perturb coastal marine ecosystem. Estuaries 10: 284-290.
Cosper, E.M., Dennison, W., Milligan, A., Carpenter, E.J., Lee, C, Holzapfel, ]., and Milanese,
L. 1989. An examination of the environmental factors important to initiating and
sustaining "brown tide" blooms, in: Cosper, E.M., Carpenter, E.J., and Bricelj, V.M.
(eds.). Novel Phytoplankton Blooms: Causes and Impacts of Recurrent Brown Tides and
Other Unusual Blooms. Lecture Notes on Coastal and Estuarine Studies. Berlin: Springer-
Verlag.
Cosper, E.M., Lee, C., and Carpenter, EJ. 1990. Novel "brown tide" bloom in Long Island
embayments: a search for the causes, in: Graneli, E., Sundsttrom, B., Edler, L., and
Anderson D.M. (eds.). Toxic Marine Phytoplankton. Elsevier, New York.
Cosper, E.M. 1991. Recent and historical novel algal blooms, Monospecific blooms occurred
along northeast coast in 1980s. 3(#2): 3-6. Waste Management Research Report (SUNY
Buffalo, SUNY, Stony Brook, and Cornell University.
Courtney, K. and Andrews, J. 1985. Neonatal and maternal blood burdens of hexachlorobenzene
(HCB) in mice: gestational exposure and lactational transfer. Fundamental and Applied
Toxicology 5(2): 265-277.
Cox, C., Clarkson, T.W., March, D.O., Amin-Zaki, L., Tikriti, S., and Myers, G.G. 1989. Dose-
response analysis of infants prenatally exposed to methyl mercury: an application of a
single compartment model to single-strand hair analysis. Environmental Research 49:
318-332.
Cranmer, J., Cranmer, M., and Goad, P. 1984. Prenatal chlordane exposure: effects on plasma
corticosterone concentrations over the lifespan of mice. Environmental Research 35(1):
204-210.
Cross, J.N., Hardy, J.T., Hose, J.E., Hershelman, G.P., Antrim, L.D., Gossett, R.W., and
Crecelius, E.A. 1987. Contaminant concentrations and toxicity of sea-surface microlayer
near Los Angeles, California. Marine Environment Research 23: 307-323.
Csaba, G., Mag, O., Inczefi-Gonda, A., and Szeberenyi, S. 1991. Persistent influence of neonatal
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) treatment on glucocorticoid receptors and
on the microsomal enzyme system. Journal of Developmental Physiology 15(6): 337-340.
135
-------
Custer, T., Weseloh, D., Stafford, C, and Braune, B. 1988. Organochlorine concentrations in
eggs of common terns from four Ontario colonies, 1981. In press.
Custer, T.W. and Peterson, D.W. 1991. Growth rates of great egret, snowy egret, and black-
crowned night heron chicks. Colonial Waterbirds 14(1): 46-50.
Cuthill, S., Wilhelmsson, A., Mason, G., Gillner, M., Pocllinger, L., and Gustafsson, J. 1988. The
dioxin receptor: a comparison with the glucocorticoid receptor. Journal of Steroid
Biochemistry 30(1-6): 277-280.
Dabeka, R., Karpinski, K., McKenzie, A., and Bajdik, C. 1986. Survey of lead, cadmium and
fluoride in human milk and correlation of levels with environmental food factors.
Foundations of Chemistry and Toxicology 24(9): 913-921.
Daly, H. 1989. Preference for unpredictable rewards occurs with high proportion of reinforced
trials or alcohol injections when rewards are not delayed. Journal of Experimental
Psycholology: animal Behavior Processes 15: 3-13.
Daly, H. 1991. Reward reduction found more aversive by rats fed environmentally contaminated
salmon. Neurotoxicology and Teratology 13: 449-453.
Daly, H. 1992a. Professor, SUNY Buffalo, New York, NY.
Daly, H. 1992b. Consumption of environmentally contaminated salmon increases work done on
a progressive ratio schedule in adult laboratory rats and their offspring. In: Isaacson, R.L.
and Jensen, K.F. (eds.). The Vulnerable Brain: Nutrition and Toxins. New York, NY:
Plenum Press. In press.
d'Argy, R., Bergman, J., and Dencker, L. 1989. Effects of immunosuppressive chemicals on
lymphoid development in foetal thymus organ cultures. Pharmacology and Toxicology 64:
33-38.
Davies, T.W., Nielsen, S.W., and Jortner, B.S. 1977. Pathology of chronic and subacute canine
methylmercurialism. Journal of the American Animal Hospital Association 13: 369-381.
Davies, D. and Mes, J. 1987. Comparison of residue levels of some Organochlorine compounds
in breast milk of the general and indigenous Canadian populations. Bulletin of
Environmental Contamination and Toxicology 39: 743-749.
Davis, C.C. 1964. Evidence for the eutrophication of Lake Erie from phytoplankton records.
Limnology & Oceanography 9: 275-283.
136
-------
Davis, D. and Safe, S. 1988. Immunosuppressive activities of polychlorinated dibenzofuran
congeners: quantitative structure-activity relationships and interactive effects. Toxicology
Applied Pharmacology 94: 141-149.
pavis, D. and Safe, S. 1989. Dose-response immunotoxicities of commercial polychlorinated
biphcnyls (PCBs) and their interaction with 23,7,8-tetrachlorodibenzo-p-dioxin.
Toxicology Letters 48: 35-43.
Dayton, L. 1991. Concern grows over toxic threats to Australia's seas. New Scientist. June 1:18.
D'Elia, C.F., Sanders, J.G., and Boynton, W.R. 1986. Nutrient enrichment studies in a coastal
plain estuary: phytoplankton growth hi large-scale, continuous cultures. Canadian Journal
of Fisheries and Aquatic Science 43: 397-406.
D'Elia, C.F. 1987. Nutrient enrichment of the Chesapeake Bay — too much of a good thing.
Environment 29: 6-33.
den Besten, PJ. 1991. Effects of cadmium and PCBs on reproduction of the sea star (Asterias
rubens). Ph.D. Thesis. University Utrecht. The Netherlands.
Dennison, W.C., Marshall, G.J., and Wigand, C. 1989. Effect of "brown tide" shading on eelgrass
(Zostera marina L.) distributions. In: Cosper, E.M., Carpenter, E.J., and Bricelj, V.M.
(eds.). Novel Phytoplankton Blooms: Causes and Impacts of Recurrent Brown Tides and
Other Unusual Blooms. Lecture Notes on Coastal and Estuarine Studies. Berlin: Springer-
Verlag.
DePinto, J.V. 1991. State of the Lake Ontario ecosystem: introduction to an ecosystem
perspective on a vital resource. Canadian Journal of Fisheries and Aquatic Science 48:
1500-1502.
DePinto, J.V., Young, T.C., and Mcllroy, L.M. 1986. Great Lakes water quality improvement.
Environmental Science and Technology 20: 752-759.
DeVito, M.J., Thomas, T., Martin, E., Umbreit, T.H., and Gallo, M.A. 1992. Antiestrogenic
action of 2,3,7,8-tetrachlorodibenzo-p-dioxin: tissue-specific regulation of estrogen
receptor in CD1 mice. Toxicology and Applied Pharmacology 113: 284-292.
Diamond, J.M. 1989. Goslings of gay geese. Nature 340: 101.
DiBartolomeis, M.J., Moore, R.W., Peterson, R.E., Christian B.J., and Jefcoate, C.R. 1987.
Altered regulation of adrenal steroidogenesis in 2,3,7,8-tetrachlorodibenzo-p-dioxin-
treated rats. Biochemical Pharmacology 36(1): 59-67.
137
-------
Dickerson, R., Howie, L., and Safe, S. 1992. The effect of 6-nitro-l,3,8-trichlorodibenzofuran
as a partial estrogen in the female rat uterus. Toxicology and Applied Pharmacology
113(1): 55-63.
Dieringer, C.S., Lamartinierc, CA., Schiller, C.M., and Lucier, G.W. 1979. Altered ontogeny of
hepatic steroid-metabolizing enzymes by pure polychlorinated biphenyl congeners.
Biochemical Pharmacology 28: 2511-2514.
Dieter, M.P. 1974. Plasma enzymes activities in coturnix quail fed graded doses of DDE,
polychlorinated biphenyl, malathion and mercuric chloride. Toxicology and Applied
Pharmacology 27: 86-98.
Dieter, M.P., Boorman, GA., Jameson, C.W., Eustis, S.L., and Uraih, L.C. 1992. Development
of renal toxicity in F344 rats gavaged with mercuric chloride for 2 weeks, or 2, 4, 6, 15,
and 24 months. Journal of Toxicology and Environmental Health 36: 319-340.
Dietrich, K.N., Krafft, K.M., Bomschein, R.L., Hammond, P.B., Berger, O., Succop, PA., and
Bier, M. 1987. Low-level fetal lead exposure effect on neurobehavioral development in
early infancy. Pediatrics 80: 721-730.
Dietz, R., Heide-Jorgensen, M.P., and Harkonen, T. 1989. Mass deaths of harbor seals (Phoca
vitulina) in Europe. Ambio 18(5): 258-264.
Doering, P.H., Oviatt, CA., Beatty, L.L., Banzon, V.F., Rice, R., Kelly, S.P., Sullivan, B.K., and
Frithsen, J.B. 1989. Structure and function in a model coastal ecosystem: silicon, the
benthos and eutrophication. Marine Ecology Progress Series 52: 287-299.
Doggett, S.M. and Rhodes, R.G. 1991. Effects of a diazinon formulation on unialgal growth rates
and phytoplankton diversity. Bulletin of Environmental Contamination and Toxicology
47: 36-42.
Doskey, P.V. and Andren, A.W. 1981a. Modeling the flux of atmospheric polychlorinated
biphenyls across the air/water interface. Environmental Science and Technology 15: 705-
711.
Doskey, P.V., and Andren, A.W. 1981b. Concentrations of airborne PCBs over Lake Michigan.
Journal of Great Lakes Research 7: 15-20.
Dougherty, R., Whitacker, M., Tang, S., Bottcher, R., Keller, M., and Kuehl, D. 1980. Sperm
density and toxic substances: a potential key to environmental health hazards. Pp. 263-
278 in: McKinney, J. (ed.). Environmental Health Chemistry. Ann Arbor, MI: Science
Publishers, Inc.
138
-------
Dynamac Corporation. Toxicological Profile for Aldrin/Dieldrin. Prepared for Agency for Toxic
Substances and Disease Registry, U.S. Public Health Service, Contract 68-D8-0056.
Eadon, G., Kraminsky, L., Silkworth, J., Aldous, K., Hilker, D., O'Keefe, P., Smith, R., Gierthy,
J., Hawley, J., Kim, N., and DeCaprio, A. 1986. Calculation of 2,3,7,8-TCDD equivalent
concentrations of complex environmental contaminant mixtures. Environmental Health
Perspective 70: 221-227.
Ebner, K., Brewster, D.W., and Matsumura, F. 1988. Effects of 2,3,7,8-tetrachlorodibenzo-p-
dioxin on serum insulin and glucose levels in the rabbit. Journal of Environmental Science
and Health B. 23(5): 427-438.
Eccles, C.U., and Annau, Z. 1987. Prenatal exposure to methyl mercury. Pp. 114-130 in: Eccles,
C.U. and Annau, Z. (eds.). The Toxicity of Methyl Mercury. Baltimore, MD: Johns
Hopkins University Press.
Edmondson, W.T. 1970. Phosphorus, nitrogen, and algae in Lake Washington after diversion of
sewage. Science 169: 690-691.
Ehrhardt, A.A. and Meyer- Bahlburg, F.L. 1981. Effects of prenatal sex hormones on gender-
related behavior. Science 211: 1312-1317.
Eisenreich, S.J. and Johnson, T.C. 1981. Airborne organic contaminants hi the Great Lakes eco
system. Environmental Science and Technology 15: 30-38.
Eisenreich, S.J. and Looney, B.B. 1982. Evidence for the atmospheric flux of PCBs to Lake
Superior. Pp. 141-156 in: MacKay, D. (ed.). Physical Behavior of PCBs in the Great
Lakes. Ann Arbor, MI: Science Publishers.
Eisenreich, S.J., and Johnson, T.C. 1983. PCBs in the Great Lakes: sources, sinks, burdens. Pp.
49-75 in: D'ltri, P.M. and Kamrin, M.A. (eds.). PCBs: Human and Environmental
Hazards. Boston, MA: Buttenvorth Publishers.
Eisenreich, S.J. and Strachan, W.MJ. 1992. Estimating Atmospheric Deposition of Toxic
Substances to the Great Lakes. Workshop Proceedings, Canada Centre for Inland Waters,
Burlington, Ontario, Canada, January 31 - February 2, 1992.
Eisenreich, S.J. and Jeremiason, J. 1992. Unpublished data. University of Minnesota.
Eisler, R. 1989 Atrazine hazards to fish, wildlife, and invertebrates: a synoptic review. U.S. Fish
and Wildlife Service; U.S. Department of the Interior. Contaminant Hazard Reviews;
Report 18. Biological Report 85(1.18).
139
-------
Elissaide, M. and Clark, D. 1979. Testosterone metabolism by hexachlorobenzene-induced
hepatic microsomal enzymes. American Journal of Veterinarian Research 40(12): 1762-
1766.
Ellenton, J.A., Brownlee, L.J., and Hollcbone, B.R. 1985. Aryl hydrocarbon hydrolylase levels
in herring gull embryos from different locations on the Great Lakes. Environmental
Toxicology & Chemistry 4: 615-622.
Eisner, H., Hodel, B., Suter, K.E., Oeklke, D., Ulbrich, B., Schreiner, G., Cuomo, V., Cagiono,
RA., Rosengren, L.E., Karlsson, J.E., and Haglid, K.G. 1988. Detection limits of different
approaches in behavioral teratology, and correlation of effects with neurochemical
parameters. Neurotoxicology and Teratology 10: 155-167.
Erdman, T. 1988. Report to US Fish and Wildlife Service on common and Roster's tern
productivity on Kidney Island confined disposal facility, Green Bay, 1987 with
supplemental necropsy and pathology reports. Green Bay: University of Wisconsin, April
1, 1988.
Erickson, J., Mulinare, J., McClain, P., Fitch, T., James, L., McCleam A., and Adams, Jr., M.
1984. Vietnam veterans' risks for fathering babies with birth defects. Journal of the
American Medical Association 242(7): 903-912.
Eriksson, P., Archer, T., and Fredriksson, A. 1990a. Altered behavior hi adult mice exposed to
a single low dose of DDT and its fatty acid conjugate as neonates. Brain Research 514(1):
141-142.
Eriksson, P., Nilsson-Hakansson, L., Nordberg, A., Aspberg. A., and Fredriksson, A. 1990b.
Neonatal exposure to DDT and its fatty acid conjugate: effects on cholinergic and
behavioural variables in the adult mouse. Neurotoxicology 11(2): 345-354.
Ernhart, C.B., Morrow-Tlucak, M., Marler, M.R., and Wolf, A.W. 1987. Low-level lead
exposure in the prenatal and early preschool periods: early preschool development.
Neurotoxicology and Teratology 9: 259-270.
Eroschenko, V.P. and Cooke, P.S. 1990. Morphological and biochemical alterations in
reproductive tracts of neonatal female mice treated with the pesticide methoxychlor.
Biology of Reproduction 42(3): 573-583.
Eroschenko, V.P. 1991. Ultrastructure of vagina and uterus in young mice after methoxychlor
exposure. Reproductive Toxicology 5(5): 427-435.
falck, Jr., F., Ricci, Jr., A., Wolff, M.S., Godbold, J., and Deckers, P. 1992. Pesticides
andpolychlorinated biphenyl residues in human breast lipids and their relation to breast
cancer. Archives of Environmental Health 47(2): 143-146.
140
-------
Falk, S.A., Klein, R., Haseman, J.K., Sanders, G.M., and Talley, F.A. 1974. Acute
methylmercury intoxication and ototoxicity in guinea pigs. Archives of Pathology 97:
297-305.
Fein, G.G., Schwartz, P.M., Jacobson, S.W., and Jacobson, J.L. 1983. Environmental toxins and
behavioral development: a new role for psychological research. American Psychologist
38: 1188-1197.
Fein, G.G., Jacobson, J., Jacobson, S., Schwartz, P., and Dowler, J. 1984. Prenatal exposure to
polychlorinated biphenyls: effects on birth size and gestional age. Journal of Pediatrics.
105: 315-320.
Fevold, H.R. 1983. Regulation of the adrenal and gonadal microsomal mixed function oxygenases
of steroid hormone biosynthesis. Annual Review of Physiology 43: 19-36.
Filippini, G., Bordo, B., Crenna, P., Massetto, N., Musicco, M., and Boeri, R. 1981. Relationship
between clinical and electrophysiological findings and indicators of heavy exposure to
2,3,7,8-tetrachlorodibenzo-dioxin. Scandinavian Journal of Work And Environmental
Health 7: 257-262.
Fimreite, N. 1971. Effects of dietary methylmercury on ringnecked pheasants. Canadian Wildlife
Service Occasional Paper No. 9.
Fine, J.S., Gasiewicz, T.A., and Silverstone, A.E. 1988. Lymphocyte stem cell alterations
following perinatal exposure to 2,3,7,8-terrachlorodibenzo-p-dioxin. Molecular
Pharmacology 35: 18-25.
Fine, J.S., Gasiewicz, T.A., and Silverstone, A.E. 1989. Lymphocyte stem cell alterations
following perinatal exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Molecular
Pharmacology 35(1): 18-25.
Fine, J.S., Gasiewicz, T.A., Fiore, N.C., and Silverstone, A.E. 1990a. Prothymocyte activity is
reduced by perinatal 2,3,7,8-tetrachlorodibenzo-p-dioxin exposure. Journal of
Pharmacology and Experimental Therapeutics 255(1): 128-32.
Fine, J.S., Silverstone, A.E., and Gasiewicz. T.A. 1990b. Impairment of prothymocyte activity
by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Journal of Immunology 144(4): 1169-1176.
Fingerhut, M.A., Halperin, W.E., Marlow, DA., Piacitelli, L.A., Honchar, P.A., Sweeney, M.H.,
Greife, A.L., Dill,. P.A., Steenland, K., and Suruda, A.J. 1991. Cancer mortality in
workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. New England Journal of
Medicine 324: 212-218.
141
-------
Fisher, N.S. 1975. Chlorinated hydrocarbon pollutants and photosynthesis of marine
phytoplankton: a reassessment. Science 189 (4201): 463-464. August 8.
Fisher, D.C. and Oppenheimer, M. 1991. Atmospheric nitrogen deposition and the Chesapeake
Bay Estuary. Ambio 20: 102-108.
Flett, RJ., Schindler, D.W., Hamilton, R.D., and Campbell, N.E.R. 1980. Nitrogen fixation in
Canadian Precambrian shield lakes. Canadian Journal of Fisheries and Aquatic Science
37: 494-505.
Fossi, C, Leonzio, C, and Focardi, S. 1986. Increase of organochlorines and MFO activity in
water birds wintering in an Italian lagoon. Bulletin of Environmental Contamination and
Toxicology 37: 538-548.
Fossi, C., Leonzio, C., and Focardi, S. 1986. Mixed function oxidase activity and cytochrome P-
450 forms in black-headed gulls feeding in different areas. Marine Pollution Bulletin 17:
546-548.
Fossi, C., Leonzio, C., and Focardi, S. 1986. Increase of organochlorines and MFO activity in
water birds wintering in an Italian lagoon. Bulletin of Environmental Contamination and
Toxicology 37: 538-548.
Fox, GA. 1992. Epidemiological and pathobiological evidence of contaminant-induced
alterations in sexual development in free-living wildlife. In: Colborn, T., and Clement,
C. (eds.). Chemically-induced Alterations in Sexual and Functional Development: The
Human-Wildlife Connection. Princeton,.NJ: Princeton Scientific Publishing, Inc. In press.
Fox, GA. and Peakali, D.B. 1991. Effects of contaminants on wildlife species. Pp. 493-755 in:
Toxic Chemicals in the Great Lakes and Associated Effects. Volume U, Effects.
Environment Canada, Department of Fisheries and Oceans, Health and Welfare Canada.
Cat. No. En 37-95,96/1990-IE.
Fox, GA. and Weseloh, D.V. 1987. Colonial waterbirds as bio-indicators of environmental
contamination in the Great Lakes. ICBP Technical Publication 6: 209-216.
Fox, G., Gilman, A., Peakali, D., and Anderka, F. 1978. Behavioral abnormalities of nesting Lake
Ontario herring gulls. Journal of Wildlife Management 42(3): 477-483.
Fox, GA., Weshloh, D.V., Kubiak, T.J., and Erdman, T.C. 1991. Reproductive outcomes in
colonial fish-eating birds: a biomarker for developmental toxins in Great Lakes food
chains. Journal of Great Lakes Research 17: 153-157.
142
-------
Frank, R., Holdrinet, M., Braun, H.E., Thomas, R.L., and Kemp, A.L.W. 1977. Organochlorine
insecticides and PCBs in sediments of Lake St. Clair (1970 and 1974) and Lake Erie
(1971). Science of the Total Environment 8: 205-227.
Frank, R., Holdrinet, M., and Suda, P. 1979. Organochlorine and mercury residues in wild
mammals in southern Ontario, Canada, 1973-1974. Bulletin of Environmental
Contamination and Toxicology 22: 500-507.
Freeman, H.C. and Sangalang, G.B. 1977. A study on the effects of methyl mercury, cadmium,
arsenic, selenium, and a PCB (Arochlor 1254) on adrenal and testicular steroidogeneses
in vitro, by the grey seal (Halichoerus gyrpus). Archives of Environmental Contamination
and Toxicology 5: 369-383.
Freeman, H.C., Sangalang, G.B., and Flemming, B. 1982. The sublethal effects of
polychlorinated biphenyl (Arochlor 1254) diet on the Atlantic cod Gadus morhua. Science
and the Total Environment 4: 1-11.
Fried, P.A. 1982. Marihuana use by pregnant women and effects on offspring: an update.
Neurobehavioral Toxicology and Teratology 4: 451-454.
Friend, M. and Trainer, D.O. 1970. Polychlorinated biphenyls: interaction with duck hepatitis
virus. Science 170: 1314-1316.
Fries, C.R. and Lee, R.G. 1984. Pollutant effects on the mixed function oxygenase (MFO) and
reproductive systems of the marine polychaete Nereis virens. Marine Biology 79: 187-
193.
Frithsen, J.B., Oviatt, CA., Pilson, M.E.Q., Howarth, R.W., and Cole, J.J. 1988. A comparison
of nitrogen vs. phosphorus limitation of production in coastal marine ecosystems. EOS
69(4): 1100.
Fry, D.M., and Toone, C.K. 1981. DDT-induced feminization of gull embryos. Science 231:
919-924.
Fry, D.M., Rosson, B., Bomardier, M., Ditto, M., MacLellan, K., and Bird, D.M. 1989.
Reproductive and behavioral effects of dicofol to progeny of exposed kestrels. Society of
Environmental Toxicology and Chemistry, Annual Meeting Poster.
Fry, D.M., Toone, C.K., Speich, S.M., and Peard, RJ. 1987. Sex ratio skew and breeding
patterns of gulls: demographic and lexicological considerations. Studies in Avian Biology
10: 26-43.
Fujii-Kuriyama, Y., Sogawa, K., Imataka H., Yasumoto, K., Kikuchi, Y., and Fujisawa-Sehara,
A. 1990. Transcriptional regulation of 3-methylcholanthrene-inducible P-450 gene
143
-------
responsible for metabolic activation of aromatic carcinogens. International Symposium of
the Princess Takamatsu Cancer Research Fund 21: 165-175.
Fukatsu, A., Brentjens, J.R., Killen, P.O., Kleinman, H.K., Martin, G.R., and Andres, GA. 1987.
Studies on the formation of glomerular immune deposits in Brown Norway rats injected
with mercuric chloride. Clinical Immunology and Immunopathology 45: 35-47.
funatsu, I., Yamashita, F., Ito, Y., Tseugama, S., Fanatsu, T., Yoshikane, T., Hayashi, T. Kato,
M., Yakushiji, M., Okamoto, G., Yamasaki, S., Arima, T., Kuno, T., Ide, H., and Ibe, I.
1972. Polychlorinated biphenyls (PCB)-induced fetopathy. Kurame Medical Journal 19:
43-51.
Funscth, E. and nback, N.G. 1992. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on blood and
spleen natural killer (NK) cell activity in the mouse. Toxicology Letters 60(3): 247-256.
Gallo, M. 1988. Rationale for a hormone-like mechanism of 2,3,7,8-TCDD for use in risk
assessment. Appendix F. In: U.S. Environmental Protection Agency. A Cancer Risk-
Specific Dose Estimate for 2,3,7,8-TCDD (Review Draft) (Appendices A through F).
Office of Health and Environmental Assessment. EPA/600/6-88/007Ab.
Gappa, J., Lopez, J., Tablado, A., and Magaldi, N.H. 1990. Influence of sewage pollution on a
rocky intertidal community dominated by the mytilid Brachidontes rodriguezi. Marine
Ecology Progress Series 63: 163-175.
Gardner, W.S., Seitzinger, S.P., and Malczyk, J.M. 1991. The effects of sea salts on the forms
of nitrogen released from estuarine and freshwater sediments: does ion pairing affect
ammonium flux? Estuaries 14: 157-166.
Gasiewicz, T.A. and Rucci, G. 1991. Alpha-naphthoflavone acts as an antagonist of 2,3,7,8-
tetra'chlorodibenzo-p-dioxin by forming an inactive complex with the Ah receptor.
Molecular Pharmacology 49(5): 607-612.
Geike, F. and Parasher, C.D. 1978. Effect of hexachlorobenzene (HCB) on photosynthetic oxygen
evolution and respiration of Chlorella pyrenoidosa. Bulletin of Environmental
Contamination and Toxicology 20: 647-651.
Gellert, RJ. Heinrichs, W.L., and Swerdloff, R.S. 1972. DDT homologues: estrogen-like effects
on the vagina, uterus, and pituitary of the rat. Endocrinology 91: 1095-1100.
Geraci, J.R. 1989. Clinical investigation of the 1987-1988 mass mortality of bottlenose dolphins
along the US central and south Atlantic coasts. Final Report, NMFS, US Navy, Office of
Naval Research.
144
-------
Giam, C.S., Wong, M.K., Hanks, A.R., and Sackett, W.M. 1973. Chlorinated hydrocarbons in
plankton from the Gulf of Mexico and Northern Caribbean. Bulletin of Environmental
Contamination and Toxicology 11: 376-382.
van Giersbergen, P., Danse, L., van Velsen, F., and van Leeuwen, F. 1984. Does b-HCH exerts
an oestrogenic effect? Verhandheling van de Faculteit Landbowwetschappen te Gent
49/3b: 1195-1202.
Gierthy, J., Lincoln, D., Gillespie, M., Seeger, J., Martinez, H., Dickerman, H., and Kumar, S.
1987. Suppression of estrogen-regulated extracellular tissue plasminogen activator activity
of MCF-7 cells by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Cancer Research 47:6198-6203.
Gieson, W.B J.T., van Katwijk, M.M., and den Hartog, C. 1990. Eelgrass condition and turbidity
in the Dutch Wadden Sea. Aquatic Botany 37: 71-85.
Giesy, J.P., Ludwig, J.P., and Kubiak, T.J. 1991. Effects of PCBs and other halogenated aromatic
hydrocarbons on Caspian tern reproduction in the upper Great Lakes (in progress).
Giesy, J.P., Jones, P.D., Tillit, D.E., Newsted, J.L., and Verbrugge, DA. 1990. Toxicity in eggs
of Great Lakes colonial waterbirds 1986-1989. Abstract and paper presented at the
International Association of great Lakes Researchers Symposium. Windsor, Canada.
Gilbertson, M. 1974a. Seasonal changes in organochlorine compounds and mercury in common
terns of Hamilton Harbour, Ontario. Bulletin of Environmental Contamination and
Toxicology 12(6): 726-732.
Gilbertson, M. 1974b. Pollutants hi breeding herring gulls. The Canadian Field-Naturalist 88(3)
273-280.
Gilbertson, M. 1975. A Great Lakes tragedy. Nature Canada 4: 22-25.
Gilbertson, M., and Fox, G. 1977. Pollutant-associated embryonic mortality of Great Lakes
herring gulls. Environmental Pollution 12: 211-216.
Gilbertson, M., Kubiak, T.J., Ludwig, J.P., and Fox, G. 1991. Great Lakes embryo mortality,
edema, and deformities syndrome (GLEMEDS) in colonial fish-eating birds: similarity
to chick edema disease. Journal of Toxicology and Environmental Health 33: 455-520.
Gilbertson, M. Secretary. 1992. Water Quality Board, International Joint Commission (IJQ.
Windsor, Ontario.
Gill, T.S., Tewari, H., and Pande, J. 1990. Use of the fish enzyme system in monitoring water
quality: effects of mercury on tissue enzymes. Comparative Biochemistry and Physiology
97: 287-292.
145
-------
Oilman, A., Beland, P., Colborn, T., Fox, G., Gicsy, J., Hesse, J., Kubiak, T., and Piekarz, D.
1991. Chapter 4. Environmental and Wildlife Toxicology of Exposure to Toxic
Chemicals. Pp. 295 in: Flint, R.W. and Vena, J. (eds.). Human Health Risks From
Chemical Exposure: The Great Lakes Ecosystem. Chelsea, MI: Lewis Publishers.
Oilman, A., Hallett, DJ., Fox, G., Allan, L., Learning, W., and Peakall, D. 1978. Effects of
injected organochlorines on naturally incubated herring gull eggs. Journal of Wildlife
Management. 42: 484-493.
Oilman, A., Peakall, D., Hallett, DJ., Fox, G., and Noistrom, R. 1977. Herring gulls (Lams
argentatus) as monitors of contamination in the Great Lakes. Pp. 280-289 in: Animals
as Monitors of Environmental Pollution. National Academy of Sciences.
Glooschenko, W.A., Strachan, W.MJ., and Sampson, R.CJ. 1976. Distribution of pesticides and
polychlorinated biphenyls in water, sediments and seston of the Upper Great Lakes-1974.
Pesticides Monitoring Journal 10: 61-67.
Golden, N.L., Sokol, RJ., Kuhnert, B.R., and Bottoms, S. 1982. Maternal alcohol use and infant
development. Pediatrics 70: 931-934.
Goldsborough, L.D. and Brown, DJ. 1988. Effect of glyphosate (roundup formulation) on
periphytic algal photosynthesis. Bulletin of Environmental Contamination and Toxicology
41: 253-260.
Gorski, J.R., Muzi, G., Wever, L.W., Pereira, D.W., and latropoulos, M.J. 1988a. Elevated
plasma corticosterone levels and histopathology of the adrenals and thymuses in 2,3,7,8-
tetrachlorodibenzo-p-dioxin-treated rates. Toxicology 53(1): 19-32.
Gorski J.R., Weber, L.W., and Rozman, K. 1988b. Tissue-specific alterations of de novo fatty
acid synthesis in 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-treated rats. Archives of
Toxicology 62(2-3): 146-151.
Gorski, J.R., Rozman, T., Greim, H., and Rozman, K. 1988c. Corticosterone modulates acute
toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in male Sprague-Dawley rats.
Fundamental and Applied Toxicology 11(3): 494-502.
Gorski, J.R., Lebofsky, M., and Rozman, K. 1988d. Corticosterone decreases toxicity of 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD) in hypophysectomized rats. Journal of Toxicology
and Environmental Health 25(3): 349-360.
Gorski, J.R. and Rozman, K. 1987. Dose-response and time course of hypothyroxinemia and
hypoinsulinemia and characterization of insulin hypersensitivity in 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD)-treated rats. Toxicology 44(3): 297-307.
146
-------
Gorski J.R., Weber, L.W., and Rozman, K. 1990. Reduced gluconeogenesis in 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD)-treated rats. Archives of Toxicology 64(1): 66-71.
Gorsline, J., Holmes, W.N., and Cronshaw, J. 1981. The effects of ingested petroleum on the
naphthalene-metabolizing properties of liver tissue in seawater-adapted Mallard duck
(Anas platyrhynchos). Environmental Research 24: 377-390.
Government of Canada. 1991. Toxic Chemicals in the Great Lakes and Associated Effects.
Volume n, Effects. Environment Canada, Department of Fisheries and Oceans, Health and
Welfare Canada. Catalogue Number EN-37-95,96/1990-lE.
Graham, M.J., Lucier, G.W., Linko, P., Maronpot, R.R. and Goldstein, J.A. 1988. Increases in
cytochrome P-450 mediated 17 beta-estradiol 2-hydroxylase activity in rat liver
microsomes after both acute administration and subchronic administration of 2,3,7,8-
tetrachorodibenzo-p-dioxin in a two-stage hepatocarcinogenesis model. Carcinogenesis
9: 1935-41.
Graneli, E. 1978. Algal assay of limiting nutrients for phytoplankton production in the Oresund.
Vatten 2: 117-128.
Graneli, E. 1981. Bioassay experiments in the Falsterbo Channel — nutrients added daily. Kieler
meeresforsch. Sonderheft 5: 82-90.
Graneli, E. 1984. Algal growth potential and limiting nutrients for phytoplankton production in
Oresund water of Baltic and Kattegat origin. Limnologica (Berlin) 15: 563-569.
Graneli, E., Wallstrom, K., Larsson, U., Graneli, W., and Elmgren, R. 1990. Nutrient limitation
of primary production in the Baltic Sea area. Ambio 19: 142-151.
Grant, D., Mes, J., and Frank, R. 1976. PCB residues in human adipose tissue and milk. In:
National Conference Proceedings on PCBs. U.S. Environmental Protection Agency. EPA
Grassle, B. and Biessmann, A. 1982. Effects of DDT, polychlorinated biphenyls and thiouracil
on circulating thyroid hormones, thyroid histology and eggshell quality in Japanese quail
(Coturnix coturnix japonica). Chemico-Biological Interactions 42: 371-377.
Gray, J.S., and Paasche, E. 1984. On marine eutrophication. Marine Pollution Bulletin 15: 349-
350.560/6-75-004.
Gray, L.E. Jr., Ostby, J.S., Ferrell, J.M., Sigmon, E.R., and Goldman, J.M. 1988. Methoxychlor
induces estrogen-like alterations of behavior and the reproductive tract in the female rat
and hamster: effects on sex behavior, running wheel activity, and uterine morphology.
Toxicology and Applied Pharmacology 96(3): 525-540.
147
-------
Great Lakes Water Quality Board. 1987. Report on Great Lakes Water Quality, International
Joint Commission, Winsdor, Ontario. Canada.
Greig, J.B., Jones, G., Butler, W.H., and Barnes, J.M. 1973. Toxic effects of 2,3,7,8-
tetrachlorodibenzo-p-dioxin. Food Cosmetology and Toxicology 11: 585-595.
Groffman, P.M. and Jaworski, N.A. 1991. Watershed nitrogen management: Upper Potomac
River basin case study. Pp. 47-59 in: Perspectives in the Chesapeake System: A Research
and Management Partnership. Baltimore, MD: Chesapeake Bay Research Consortium
Publication #137.
Guengerich, P.P. 1991. Reactions and significance of cytochrome P-450 enzymes. Journal of
Biological Chemistry 266(16): 10019-10022.
Guengerich, P.P. 1992. Human cytochrome P-450 enzymes. Life Sciences 50: 1471-1478.
Guery, J.C., Druet, E., Glotz, D., Hirsch, R, Mandet, C, De-Heer, R, and Druet, P. 1990.
Specificity and cross-reactive idiotypes of anti-glomerular basement membrane
autoantibodies in HgC12-induced antoimmune glomerulonephritis. European Journal of
Immunology 20: 93-100.
Gustafsson, J.A., Mode, A., Norstedt, G., and Skett, P. 1983. Sex steroid induced changes in
hepatic enzymes. Annual Review of Physiology 45: 51-60.
Gutkina, N. and Mishin, V. 1986. Immunochemical evidences that hexachlorobenzene induces
two forms of cytochrome p-450 in the rat liver chromosomes. Chemical and Biological
Interactions 58(1): 57-68.
Haake, J., Kelley, M., Keys, B., and Safe, S. 1987. The effects of organochlorine pesticides as
inducers of testosterone and benzo[a]pyrene hydroxylases. General Pharmacology 18(2):
165-169.
Haegele, MA. and Tucker, R.K. 1974. Effects of 15 common environmental pollutants on
eggshell thickness in mallards and coturnix. Bulletin of Environmental Contamination and
Toxicology 11: 09-102.
Hall, L.W., Jr., Hall, W.S., Bushong, S.J., and Herman, R.L. 1987. In situ striped bass (Morone
saxitilis) contaminant and water quality studies in the Potomac River. Aquatic Toxicology
10: 73-99.
Hall, L.W., Jr., Bushong, S.J., Zigenfuss, M.C., and Hall, W.S. 1988a. Concurrent mobile on-site
and in situ striped bass contaminant and water quality studies in the Choptank River and
Upper Chesapeake Bay. Environmental Toxicology and Chemistry 7: 815-830.
148
-------
Hall, L.W., Jr., Zigenfuss, M.C., Bushong, S.J., and Unger, M.A. 19885. Striped bass
contaminant and water quality studies in the Potomac River and Upper Chesapeake Bay -
- annual contaminant and water quality evaluations in east coast striped bass habitats.
Report, Johns Hopkins University, Applied Physics Laboratory, Shadyside, Maryland.
Hallegraeff, G.M., Steffensen, DA., and Wetherbee, R. 1988. Three esruarine dinoflagellates that
can produce paralytic shellfish toxins. Journal of Plankton Research 10: 533-541.
Hallett, D., Norstrom, R., Onuska, F., and Comba, M. 1982. Incidence of chlorinated benzenes
and chlorinated ethylenes in Lake Ontario herring gulls. Chemosphere 11(3): 277-285.
Hamilton, P.C., Jackson, G.S., Kaushik, N.K., and Solomon, K.R. 1987. The impact of atrazine
on lake periphyton communities, including carbon uptake dynamics using track
autoradiography. Environmental Pollution 46: 83-103.
Hans, S.L., Marcus, J., Jeremy, R.J., and Auerbach, J.G. 1984. Neurobehavioral development of
children exposed in utero to opioid drugs. In: Neurobehavioral Teratology. Amsterdam:
Elsevier.
Hansson, S. and Rudstam, L.G. 1990. Eutrophication and Baltic fish communities. Ambio 19:
123-125.
Harada, M. 1976. Intrauterine poisoning: clinical and epidemiological studies and significance
of the problem. Bulletin of the Institute of Constitutional Medicine. Kumamoto
University. 25 (Supp.): 1-69.
Harada, M. 1977. Congenital alkyl mercury poisoning (Congenital Minamata Disease).
Paediatrician 6: 58-68.
Harding, L.W., Jr. 1976. Polychlorinated biphenyl inhibition of marine phytoplankton
photosynthesis in the northern Adriatic Sea. Bulletin of Environmental Contamination and
Toxicology 16(5): 559-566.
Hargis, WJ., Jr. and Zweraer, D.E. 1988. Some histological gill lesions of several estuarine
finfishes related to exposure to contaminated sediments: a preliminary report. Pp. 474-487
in: Understanding the estuary: Advances in Chesapeake Bay research. Proceedings of a
Conference, 29-31 March 1988. Baltimore, Maryland. Chesapeake Research Consortium
Publication Number 129. CBP/TRS 24/88.
Hargrave, B.T, Harding, G.C., Voss, W.P., Erickson, P.E., Fowler, B.R., and Scott, V. 1992.
Organochlorine pesticides and polychlorinated biphenyls in the Arctic Ocean food web.
Archive of Environmental Contamination and Toxicology 22: 41-54.
149
-------
Harris, HJ. 1988. Persistent toxic substances and birds and mammals in the Great Lakes. Pp.
557-569 in: Evans, M.S. (ed.). Toxic Contaminant and Ecosystem Health: A Great Lakes
Focus. New York, NY: John Wiley & Sons.
Harris, HJ. 1990. Marshes, Forster's tern, and microcontaminants in Green Bay. Paper presented
at preserving Great Lakes Wetlands: an environmental agenda. Conference sponsored by
the Great Lakes Wetlands Policy Consortium. Buffalo, NY. May 15.
Harris, M., Zacharewski, T, and Safe, S. 1990. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin
and related compounds on the occupied nuclear estrogen receptor in MCF-7 human breast
cancer cells. Cancer Research 50(12): 3579-3584.
•
Harwood, J., Carter, S.D., Hughes, D.E., Bell, S.C., Baker, J.R., and Cornwall, HJ. 1989. Seal
disease predictions. Nature 339: 670.
Haseltine, S.D., Heinz, G., Reichel, W., and Moore, J. 1981. Organochlorine and metal residues
in eggs of waterfowl nesting on islands in Lake Michigan off Door County, Wisconsin,
1977-78. Pesticide Monitoring Journal 15(2): 90-97.
Haseltine, S.D., Peterle, T.J., and Nagode, L. 1981. Physiology of the eggshell thinning response
to DDE. Transactions of the International Congress of Game Biology 12: 237-243.
Hayes, M., Smith, L, Crane, T., Kocal, T., Hicks, B., and Ferguson, H. 1987. Pathogenesis of
skin and liver neoplasms in white suckers (Catostomus commersoni) from polluted areas
in Lake Ontario. Abstract and Presentation of Fourteenth Aquatic Toxicity Workshop,
Toronto. November, 1987.
Heaton, S.N., Aulerich, RJ., and Bursian, S.J. 1991. Reproductive Effects of feeding Saginaw
Bay source fish to Ranch mink. Pp. 24-25 in: Schneider, S. and Campbell, R. (eds.).
Cause-Effect Linkages II Symposium Abstracts. Michigan Audubon Society, September
27-28, 1991.
Hecky, P.E. and Kilham, P. 1988. Nutrient limitation of phytoplankton in freshwater and marine
environments: a review of recent evidence on the effects of enrichment. Limnology &
Oceanography 33: 796-822.
Hedge, GA., Colby, H.D., and Goodman, R.L. 1987. Clinical Endocrine Physiology.
Philadelphia, PA: W.B. Saunders, Co.
Helder, T. 1980. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on early life stages of
the pike (Esox lucius L.). Science of the Total Environment 14: 255-264.
Helle, E., Olsson, M., and Jensen, S. 1976a. DDT and PCB levels and reproduction in ringed
seals from the Bothnian Bay. Ambio 5(5-6): 261-263.
150
-------
Helle, E., Olsson, M., and Jensen, S. 1976b. PCB levels correlated with pathological changes in
seal uteri. Ambio 5(4): 188-189.
Hellou, J. and Payne, J.F. 1986. Effect of petroleum hydrocarbons on the biliary bile acid
composition of rainbow trout (Salmo gairdneri). Comparative Biochemistry and
Physiology 84C: 257-261.
Helz, G.R. and Huggett, RJ. 1987. Contaminants in Chesapeake Bay: the regional perspective.
Pp. 270-297 in: Majumdar, S.K., Hall, Jr., L.W., and Austin, H.M. (eds.). Contaminant
problems and management of living Chesapeake Bay resources. Philadelphia, PA:
Academy of Sciences.
Henry, E.C. and Gasiewicz, TA. 1987. Changes in thyroid hormones and thyroxine
glucuronidation in hamsters compared with rats following treatment with 2,3,7,8-
tetrachorodibenzo-p-dioxin. Toxicology and Applied Pharmacology 89(2): 165-174.
Herigstad, R.R., Whitehair, C.K. Beyer, N., Mickelsen, O., and Zabik, M.J. 1972. Chronic
methylmercury toxicosis in calves. Journal of the American Veterinary Medical
Association. 160: 173-182.
Hersh, C.M. and Crumpton, W.G. 1987. Determination of growth rate depression of some green
algae by Atrazine. Bulletin of Environmental Contamination and Toxicology 39: 1041-
1048.
Hertzler, D.R. 1990. Neurotoxic behavioral effects of Lake Ontario salmon diets in rats.
Neurotoxicology and Teratatology 12: 139-143.
Higuchi, K. (ed.). 1976. PCB poisoning and pollution. New York, NY: Academic Press.
Hines, M. 1982. Prenatal gonadal hormones and sex differences in human behavior. Psychology
Bulletin 92: 56-80.
Hoagland, K.D. and Drenner, R.W. 1991. Freshwater community responses to mixtures of
agricultural pesticides: synergistic effects of Atrazine and Bifenthrin. Texas Water
Resources Institute. Texas A&M University Technical Report No. 151. April.
Hodgins, H.O., Gronland, W.D., Mighell, H.L., Hawkes, J.W., and Robisch, P.A. 1977. Effect
of crude oil on trout reproduction. Pp. 143-150 in: Wolfe, DA., Anderson, J.W., Button,
D.K., Malins, D.C.,.Roubal, T., and Varanasi, U. (eds.). Fate and Effects of Petroleum
Hydrocarbons in Marine Ecosystems and Organisms. New York, NY: Pergamon Press.
Hodson, P.V.H., Ralph, K.M., Gray, B., and McWhirter, M. 1989. Mixed function oxidase
(MFO) activity of Great Lakes fish. Poster Session at the Tenth Annual Meeting of the
151
-------
Society of Environmental Toxicology and Chemistry, October 28 to November 2,
Toronto, Ontario. Canada.
Hoff, R.M., Muir, D.C.G., and Grift, N.P. 1992a. Annual cycle of polychlorinated biphenyls and
organochlorinc pesticides in air. in Southern Ontario. I. Air concentration data.
Environmental Science and Technology 26: 266-275.
Hoff, R.M., Muir, D.C.G., and Grift, N.P. 1992b. Annual cycle of polychlorinated biphenyls and
organohalogen pesticides in air in Southern Ontario, n. Atmospheric transport and
sources. Environmental Science and Technology 16: 276-283.
Hoffman, DJ., Rattner, BA., Sileo, L., Docherty, D., and Kubiak, TJ. 1987. Embryotoxicity,
teratogenicity and aryl hydrocarbon hydroxylase activity in Forster's Terns on Green Bay,
Lake Michigan. Environmental Research 42: 176-184.
Hoffman, DJ., Rattner, BA., Bunck, C.M., Krynitsky, A., Ohlendorf, H.M., and Lowe, R.W.
1986. -Association, between PCBs and lower embryonic weight in black-crowned night
herons in San Francisco Bay. Journal of Toxicology and Environmental Health 19: 383-
391.
Holsapple, M.P., McNerney, P.J., Barnes, D.W., and White, KX. 1984. Suppression of humoral
antibody production by exposure to 1,2,3,6,7,8-hexachlorodibenzo-dioxin. Journal of
Pharmacology and Experimental Therapy 231: 518-526.
Holsapple, M.P., Snyder, N.K., Wood, S.C., and Morris, D.L. 1991. A review of 2,3,7,8-
tetrachorodibenzo-p-dioxin-induced changes in immunocompetence: 1991 update.
Toxicology 69(3): 219-255.
Hose, J.E., Cross, J.N., Smith, S.G., and Diehl, D. 1989. Reproductive impairment in a fish
inhabiting a contaminated coastal environment off of Southern California. Environmental
Pollution 57: 139-148.
House, R.V., Lauer, L.D., Murray, M.J., Thomas, P.T., Ehrlich, J.P., Burleson, G.R., and Dean,
J.H. 1990. Examination of immune parameters and host resistance mechansisms in
B6C3F1 mice following adult exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Journal
of Toxicology and Environmental Health 31: 203-215.
Howard, J.D. and Mottet, N.K. 1986. Effects of methylmercury on the morphogenesis of the rat
cerebellum. Teratology 34: 89-95.
Howarth, R.W. and Cole, J.J. 1985. Molybdenum availability, nitrogen limitation, and
phytoplankton growth in natural waters. Science 229: 653-655.
152
-------
Howarth, R.W. 1988. Nutrient limitation of net primary production in marine ecosystems. Annual
Review of Ecology & Systematics 19: 89-110.
Howarth, R.W., Marino, R., Lane, J., and Cole, JJ. 1988a. Nitrogen fixation in freshwater,
estuarine, and marine ecosystems. 1. Rates and importance. Limnology & Oceanography
33: 669-687.
Howarth, R.W., Marino, R., and Cole, J.J. 1988b. Nitrogen fixation in freshwater, estuarine, and
marine ecosystems. 2. Biogeochemical controls. Limnology & Oceanography 33: 688-
701.
Howarth, R.W. 1991. Comparative responses of aquatic ecosystems to toxic chemical stress. Pp.
169-195 in: Col, J., Lovett, G., and Findlay, S. (eds.). Comparative Analyses of
Ecosystems: Pattersn, Mechanisms, and Theories. New York, NY: Springer-Verlag.
Howarth, R.W., Marino, R., and Cole, J.J. 1993. Why so little planktonic nitrogen fixation in
coastal marine ecosystems? Appropriate hypotheses and appropriate tests. Limnology &
Oceanography. In press.
Howie, L., Dickerson, R., Davis, D., and Safe, S. 1990. Immunosuppressive and monooxygenase
induction activities of polychlorinated diphenyl ether congeners in C57BL/6N mice:
quantitative structure-activity relationships. Toxicology and Applied Pharmacology
105(2): 254-263.
Huggett, R.J., Benser, M.E., and Unger, M.A. 1987. Polynuclear aromatic hydrocarbons in the
Elizabeth River, Virginia. Pp. 327-341 in: Dickson, K.L., Maki, A.W., and Brungs, W.A.
(eds.). Fate and effects of sediment-bound chemicals in aquatic ecosystems. Elmsford:
Pergammon Press.
Hultman, P. and Enestrom, S. 1992. Dose-response studies in murine mercury-induce
autoimmunity and immune-complex disease. Toxicology and Applied Pharmacology 113:
199-208.
Humphrey, H. 1985. Chemical contaminants hi the Great Lakes: the human health aspect.
Advances in Environmental Science and Technology. Symposium on Persistent Toxic
Substances. Minneapolis, MN: Wiley Publishers.
Hunt, Jr., G.L., Wingfield, J.C., Newman, A., and Farner, D.S. 1980. Sex ratio of western gulls
(Larus occidentalis) in southern California. Auk 97: 473-479.
Husain, M.M., Kumar, A., and Mukhtar, H. 1982. Inhibition of tissue aryl hydrocarbon
(benzo[a]pyrene) hydroxylase by 7,8-benzoflavone hi birds. Xenobiotica 12: 375-380.
153
-------
Hutzinger, O., Blumich, M., Berg, M.v.d., and Olie, K. 1985. Sources and fate of PCDDs and
PCDFs: an overview. Chemosphere 14: 581.
IARC 1986. In: O'Neill, Schuller, and Fishbein (eds.). Environmental Carcinogens Selected
Methods of Analysis. 8(71).
Dback, N.G. 1991. Effects of methyl mercury exposure on spleen and blood natural killer (NK)
cell activity in the mouse. Toxicology 67: 117-124.
flback, N.G., Sundberg, H., and Oskarsson, A. 1991. Methyl mercury exposure via placenta and
milk impairs natural killer (NK) cell function in newborn rats. Toxicology Letters 58:
149-158.
Immura, N., Miura, K., Inokawa, M., and Nakada, S. 1980. Mechanism of methylmercury
cytotoxicity: by biochemical and morpholgocial experiments using cultured cells.
Toxicology 17: 241-254.
Ingebrightsen. K., Hektoen, H., Andersson, T., Bergman, A., and Brandt. I. 1990. Species-
specific accumulation of the polychlorinated biphenyl (PCB) 2,3,3',4,4',-
pentachlorobiphenyl in fish brain: a comparison between cod (Gadus morhud) and
rainbow trout (Oncorhynchus myfdss). Pharmacology and Toxicology 67(4): 344-345.
Ingebrightsen, K., Hektoen, H., Brevik, E.M., and Oehme, M. 1991. Species-specific
accumulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the brain of cod (Gadus morhua).
Acta Veterinaria Scandinavica Supplementum 87: 309-310.
International Joint Commission (IJQ, Great Lakes Science Advisory Board. Summary report of
the workshop on Great Lakes atmospheric deposition. Windsor, Ontario. Canada. October,
1987.
International Joint Commission (IJC). 1988. Emerging issues—ongoing and emerging. Appendix
B. Bald eagle, mink, and otter chapter (draft). IJC Report.
Ireland, J.S. Mukku, V.R., Robison, A.K., and Stancel, G.M. 1980. Stimulation of uterine
deoxyribonucleic acid synthesis by l,l,l-trichloro-2-(p-chlorophenyl)-2(o-
chlorophenyl)ethane (o,p'DDT). Biochemical Pharmacology 29: 1469-1479.
Ivans, LA., Loser, E., Rinke, M., Schmidt, U., and Neupert, M. 1992. Toxicity of 2,3,7,8-
tetrachlorodibenzo-p-dioxin in rats after single oral administration. Toxicology 73(1):
53-69.
Jacobs, J.M., Carmichael, N., and Cavanagh, J.B. 1977. Ultrastructural changes in the nervous
system of rabbits poisoned with methylmercury. Toxicology and Applied Pharmacology
39: 249-261.
154
-------
Jacobs, AJ., Maniscalco, W.M., and Finkelstein, J.N. 1986. Effects of methylmercuric chloride,
cyclohexamide and colchicine on the reaggregation of dissociated mouse cerebellar cells.
Toxicology and Applied Pharmacology 86: 362-371.
Jacobson, S., Jacobson, J., Schwartz, P., and Fein, G. 1983. Intrauterine exposure of human
newborns to PCBs: Measures and exposure. Pp. 311-343 in: D'ltri, P.M., and Kamrin,
MA. (eds.). PCBs Human and Environmental Hazards. Boston, MA: Butterworth
Publishers.
Jacobson, S.W., Fein, G.G., Jacobson, J.L, Schwartz, P.M., and Dowler, J.K. 1985. The effect
of intrauterine PCB exposure on visual recognition memory. Child Development 56: 853-
860.
Jacobson, J.L. and Jacobson, S.W. 1988. New methodologies for assessing the effects of prenatal
toxic exposure on cognitive functioning in humans. In: Evans, M. (ed.). Toxic
Contaminants and Ecosystem Health: A Great Lakes Focus, Volume 21. Wiley Series.
Jacobson, J.L., Jacobson, S.W., and Humphrey, H.E.B. 1989. Effects of in utero exposure to
polychlorinated biphenyls and related contaminants on cognitive functioning in young
children. The Journal of Pediatrics. 116(1): 38-45.
Jacobson, J.L., Humphrey, H.E.B., Jacobson, S.W., Schantz, S.L., Mullin, M.D., and Welch, R.
1989. Determinants of polychlorinated biphenyls (PCBs), polybrominated biphenyls
(PBBs) and dichlorodiphenyl trichloroethane (DDT) levels in the sera of young children.
American Journal of Public Health. In press.
Jacobson, J.L., Jacobson, S.W., and Humphrey, H.E.B. 1990. Effects of exposure to PCBs and
related compounds on growth and activity in children. Neurotoxicology and Teratology
12: 319-326.
Jacobson, J.L. and Jacobson, S.W. 1991. Follow-up on children from the Michigan fish-eaters
cohort study: performance at age 4. Pp. 34-35 in: Schneider, S. and Campbell, R. (eds.).
Cause-Effects Linkages II Symposium Abstracts. Michigan Audubon Society. Lansing,
MI.
Jacobson, J.L., Jacobson, S.W., Padgett, RJ., Brumitt, G.A., and Billings, R.L. 1992. Effects of
prenatal PCB exposure on cognitive processing efficiency and sustained action.
Developmental Psychology 2892: 297-306.
Jaworski, N.B. 1981. Sources of nutrients and the scale of eutrophication problems in estuaries.
In: Neilson, B.J. and Cronin, L.E. (eds.). Estuaries and Nutrients. Humana, NY.
Jaworski, N.B., Groffman, P.M., Keller, A-A., and Prager, J.C. 1992. A watershed nitrogen and
phosphorus balance: The upper Potomac River basin. Estuaries 15: 83-95.
155
-------
Jefcoate, C.R., DiBartolomeis, M.J., Williams, C.A., and McNamara, B.C. 1987. ACTH
regulation of cholesterol movement in isolated adrenal cells. Journal of Steroid
Biochemistry 27(4-6): 721-729.
Jefferies, DJ. 1975. The role of the thyroid hi the production of sublethal effects by
organochlorine insecticides and polychlorinated biphenyl. Pp. 131-230 in: Moriarty, F.
(ed.). Organochlorine Insecticides. Persistent Organic Pollutants. New York, NY:
Academic Press.
Jennings, A.M., Wild, G., Ward, J.D., Ward. A.M. 1988. Immunological abnormalities seventeen
years after accidental exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. British Journal
Industrial Medicine 45: 701-704.
Jensen, S., Renberg, L., and Olsson, M. 1972. PCB contamination from boat bottom paint and
levels of PCB in plankton outside a polluted area. Nature 240: 358-360. December 8.
Jensen, S., Kihlstroem, J.E., Olsson, M., Lundberg, C, and Orberg, J. 1977. Effects of PCB and
DDT on mink (mustela vison) during the reproductive season. Ambio 6: 239.
Jensen, L.M., Sand-Jensen, K., Marcher, S., and Hansen, M. 1990. Plankton community
respiration along a nutrient gradient in a shallow Danish estuary. Marine Ecology Progress
Series 61: 75-85.
Jensen, A A. and Slorach, S.A. 1991. Chemical contaminants in human milk. Boston: CRC Press.
Johansson, B. 1987. Lack of effects of polychlorinated biphenyls on testosterone synthesis in
mice. Pharmacology & Toxicology 61(4): 220-223.
Johnson, K., Kaminski, N., and Munson, A. 1987. Direct suppression of cultured spleen cell
responses by chlordane and the basis for differential effects on in vivo and in vitro
immunocompetence. Journal of Toxicology and Environmental Health 22(4): 497-515.
Johnson, E. 1992. Human exposure to 2,3,7,8-TCDD and risk of cancer. Critical Review of
Toxicology 21: 451-463.
Johnson, L.L., Casillas, E., Collier, T.K., McCain, B.B., and Varanski, U. 1988. Contaminant
effects on ovarian development in English sole (Parophrys vetulus) from Puget Sound,
Washington. Canada Journal of Fisheries and Aquatic Science 45: 2133-2146.
Jones, K.L., Smith, D.W., Ulleland, C.N., and Streissguth, A.P. 1973. Pattern of malformation
in offspring'of chronic alcoholic mothers. Lancet 1: 1267-1271.
Jones, R. and Chelsky, M. 1986. Further discussion concerning porphyria cutanea tarda and
TCDD-exposure. Archives of Environmental Health 41(2): 100-103.
156
-------
Jones, M.K., Weisenburger, W.P., Sipes, I.G., and Russell, D.H. 1987. Orcadian alterations in
prolactin, corticosterone, and thyroid hormone levels and down-regulation of prolactin
receptor activity by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicology and Applied
Pharmacology 87(2): 337-350.
Jones, S.N., Jones, P.O., Ibarguen, H., Caskey, C.T., and Craigen, WJ. 1991. Induction of the
Cypla-1 dioxin-responsive enhancer in transgenic mice. Nucleic Acids Research 19(23):
6547-6551.
Kahn, A.T. and Weis, J.S. 1987. Toxic effects for mercuric chloride on sperm and egg viability
of two populations of mummichog (Fundulus heteroclitus). Environmental Pollution 48:
263-273.
Kaminski, N., Wells, D., Dauterman, W., Roberts, J., and Guthrie, F. 1986. Macrophage uptake
of a lipoprotein-sequestered toxicant: a potential route of immunotoxicity. Toxicology and
Applied Pharmacology 82(3): 474-480.
Ramp-Nielsen, L. 1974. Mud-water exchange of phosphorus and other ions in undisturbed
sediment cores and factors affecting the exchange rate. Archiv fur Hydrobiologie 13:218-
237.
Kanja, L., Skare, J., Maitai, C, and Lokken, P. 1986. Organochlorine pesticides in human milk
from different areas of Kenya 1983-1985. Journal of Toxicology and Environmental
Health 19(4): 449-464.
Kannan, N., Tanabe, S., and Tatsukawa, R. 1988. Toxic potential of non-ortho and mono-ortho
coplanar PCBs in commercial PCB preparations: "2,3,7,8-T4CDD toxicity equivalence
factors approach." Bulletin of Environmental Contaminants and Toxicology 41: 267-276.
Kapoor, I.P., Mukku, V.R., Robinson, A.K., and Stancel, G.M. 1970. Comparative metabolism
of methoxychlor, methiochlor, and DDT in mouse, insects and in a model ecosystem.
Journal of Agricultural and Food Chemistry 18: 1145-1152.
Kaye, A.M., Icekson, L, and JJndner, H.R. 1971. Stimulation by estrogens of ornithine and S-
adenosylmethionine decarboxylases in the immature rat uterus. Biochimica et Biophysica
Acta 252: 150-159.
Kazantzis, G., L*m, T., and Sullivan, K. 1988. Mortality of cadmium-exposed workers — a
five-year update. Scandinavian Journal of Work and Environmental Health. 14: 220-223.
Keesey, R.E., Boyle, P.C., Kemnitz, J.W., and Mitchel, J.S. 1976. The role of the lateral
hypothalamus in determining the body weight set point. Pp. 243-255 in: Novin, D.,
Wyrwicks, W., and Bray, G. (eds.). Hunger: Basic Mechanisms and Clinical Implications.
New York}> NY: Raven Press.
157
-------
Keith, J. 1966. Reproduction in a population of herring gulls (Larus argentatus) contaminated
by DDT. Journal of Applied Ecology 3: 57-70.
Kelly, J. and Levin, S. 1986. A comparison of aquatic and terrestrial nutrient cycling and
production processes in natural ecosystems, with reference to ecological concepts of
relevance to some waste disposal issues. In: Kullenber, G. (ed.). The Role of Oceans as
a Waste Disposal Option. Reidel, Amsterdam.
Kemp, W.M., Twilley, R.R., Stevenson, J.C., Boynton, W.R., and Means, J.C. 1983. The decline
of submerged vascular plants in upper Chesapeake Bay: summary of results concerning
possible causes. Journal of Marine Technology Society 17: 78-85.
Kerkvliet, N.I. and Baecher-Steppan, L. 1988. Suppression of allograft immunity by
3,4,53',4'^'-hexachlorobiphenyl. I. Effects of exposure on tumor rejection and cytotoxic
T cell activity in vivo. Immunopharmacology 16(1): 1-12.
Kerkvliet, N.I., Baecher-Steppan, L., Smith, B.B., and Youngberg, J.A. 1990a. Role of the Ah
locus in suppression of cytotoxic T lymphocyte activity by halogenated aromatic
hydrocarbons (PCBs and TCDD): structure-activity relationships and effects in C57B1/6
mice congenic at the Ah locus. Fundamental and Applied Toxicology 14(3): 532-541.
Kerkvliet, N.I., Steppan, L.B., Brauner, J.A., Deyo, J.A., Henderson, M.C., Tomar, R.S., and
Buhler, D.R. 19905. Influence of the Ah locus on the humoral immunotoxicity of 2,3,7,8-
tetrachlorodibenzo-p-dioxin: evidence for Ah-receptor-dependent and Ah-receptor-
independent mechanisms of immunosuppression. Toxicology and Applied Pharmacology
105(1): 26-36.
Kerper, L.E., Ballatori, N., and Clarkson, T.W. 1992. Methylmercury transport across the blood-
brain barrier by an amino acid carrier. Americal Journal of Physiology 262(5 pt 2): R761-
765.
Keys, B., Piskorska-Pliszczynska, J., and Safe, S. 1986. Polychlorinated dibenzofurans as
2,3,7,8-TCDD antagonists: in vitro inhibition of monooxygenase induction. Toxicological
Letters 31: 151.
Khalid, R.A., Patrick, W.H., and DeLaune, R.D. 1977. Phosphorus sorption characteristics of
flooded soils. Soil Science Society of American Journal 41: 305.
Khan, M.A.Q. 1984. Induction of drug-metabolizing enzymes. Pp. 129-222 in: Matsumura, F.
(ed.). Differential Toxicities of Insecticides and Halogenated Aromatics. Oxford, U.K.:
Pergamon Press.
Kirkman, R.H. 1976. A review of the literature on seagrass related to its decline in Moreton Bay,
Old. CSIRO Report Number 64.
158
-------
Kiyohara, C, Omura, M., and Hirohata, T. 1991. In vitro effects of L-ascorbic acid (vitamin C)
on aryl hydrocarbon hydroxylase activity in hepatic microsomes of mice. Mutation
Research 251(2): 227-232.
Klaunig, J. and Ruch, R. 1987a. Strain and species effects on the inhibition of hepatocytc
intercellular communication by liver tumor promoters. Cancer Letters 36: 161-168.
Klaunig, J. and Ruch, R. 1987b. Role of cyclic AMP in the inhibition of mouse hepatocyte
intercellular communication by liver tumor promoters. Toxicology and Applied
Pharmacology 91: 159-170.
Kluythmans, J.H., Brands, F., and Zandee, D.I. 1988. Interactions of cadmium with the
reporductive cycle of Mytilus edulis L Marine Environment Research 24: 198-192.
Knight, G.C. and Walker, C.H. 1982. A study of the hepatic microsomal monooxygenase of sea
birds and its relationship to organochlorine pollutants. Comprehensive Biochemistry and
Physiology 73(C): 211-221.
Knoflach, P., Albini, B., and Weiser, M.M. 1986. Autoimmune disease induced by oral
administration of mercuric chloride in Brown-Norway rats. Toxicology and Pathology
14(2): 188-193.
Kobayashi, H., Yuyama, A., Matsusaka, N., Takeno, K., and Yanagiva, I. 1979. Effects of
methylmercury chloride on various cholinergic parameters in vitro. Journal of
Toxicological Science 4: 351-362.
Kobayashi, H., Yuyama, A., Matsusaka, N., Takeno, K., and Yanagiva, I. 1981.
Neuropharmacological effect of methylmercury in mice with special reference to the
central cholinergic system. Japanese Journal of Pharmacology 31: 711-718.
Kociba, RJ., Keeler, P.A., Park, C.N., and Gehring, PJ. 1976. 2,3,7,8-Tetrachlorodibenzo-p-
dioxin (TCDD): results of a 13-week oral toxicity study in rats. Toxicology and Applied
Pharmacology 35: 553-574.
•
Kodama, H., and Ota, H. 1980. Transfer of polychlorinated biphenyls to infants from their
mothers. Archives of Environmental Health 35: 95-100.
Kodavanti. P.R., Mehrotra, B.D., Chetty. S.C., and Desaiah, D. 1988. Effect of selected
insecticides on rat brain synaptosomal adenylate cyclase and phosphodiesterase. Journal
of Toxicology and Environmental Health 25(2): 207-215.
Kolaja, G.J., and Hinton, D.E. 1979. DDT-induced reduction in eggshell thickness, weight, and
calcium is accompanied by calcium ATPase inhibition. Pp. 309-318 in: Animals as
monitor of pollutants. Washington, DC: National Academy of Sciences.
159
-------
Komulainen, H., and Tuomisto, J. 1987. The neurochemical effects of methyl mercury in the
brain. Pp. 172-188 in: Eccles, C.U. and Annau, Z. (eds.). The Taxicity of Methyl Mercury.
Baltimore, MA: Johns Hopkins University Press.
Korach, K.S., Sarver, P., Chae, K., McLachlan, J.A., and McKinney, J.D. 1988. Estrogen
receptor-binding activity of polychlorinated hydroxybiphenyls: conformationally restricted
structural probes. Molecular Pharmacology 33(1); 120-126.
Korpela, H., Loueniva, R., Yrjanheikki, £., and Kauppila, A. 1986. Lead and cadmium
concentrations in maternal and umbilical cord blood, amniotic fluid, placenta, and
amniotic membranes. American Journal of Obstetrics and Gynecology 155(5): 1086-1089.
Kovacs, K.M. and Ryder, J.P. 1981. Nest-site tenacity and male fidelity in female-female pairs
of ring-billed gulls. The Auk 98: 625-627.
Koval, P.J., Peterle, T.J., and Harder, J.D. 1987. Effects of polychlorinated biphenyls on
mourning dove reproduction and circulation progesterone levels. Bulletin of
Environmental Contamination and Toxicology 39(4): 663-670.
Krciss, K., Zack, M.M., Kimbrough, R.D., Needham, L.L., Smrek, A.L., and Jones, B.T. 1981.
Cross-sectional study of a community with exceptional exposure to DDT. Journal of the
American Medical Association 245: 1926-1930.
Kubiak, T.J. 1988. Statement on the impact of diffuse sources of toxic substances on Great Lakes
water quality. Testimony before the Subcommittee on Investigations and Oversight of the
Committee on Public Works and Transportation of the U.S. House of Representatives.
U.S. Government Printing Office Document 85-374.
Kubiak, TJ., Hams, HJ., Smith, L.M., Schwartz, T.R., Stalling, D.L., Trick, J.A., Sileo, L.,
Docherty, D.E., and Erdman, T.C. 1989. Microcontaminants and reproductive impairment
of the Forster's tern on Green Bay, Lake Michigan-1983. Archives of Environmental
Contamination and Toxicology 18: 706-727.
Kubiak, TJ. and. Best, DA. 1991. Wildlife risks associated with passageof contaminated
anadromous fish at Federal Energy Regulatory Commission Licensed Dams in Michigan.
Contaminants Program Division of Ecological Services. East Lansing, MI. August 16,
1991.
Kubiak, T. and Harris, H. 1985. Microcontaminants and reproductive impairment of the Forster's
tern on Green Bay, Lake Michigan, Final report to USFWS. September.
Kuehl, D.W., Cook, P.M., and Batterman, A.P. 1985. Studies on the bioavailability of 2,3,7,8-
TCDD from municipal incinerator fly ash to freshwater fish. Chemosphere 14: 871-872.
160
-------
Kuehl, D.W., Haebler, R., and Potter, C. 1991. Chemical residues in dolphins from the ILS.
Atlantic coast including Atlantic bottlenose obtained during the 1987/88 mass mortality.
Chemosphere 22(11): 1071-1084.
Kuhnert, B., Kuhnert, P., Debanne, S., and Williams, T. 1987. The relationship between
cadmium, zinc, and birth weight in women who smoke. American Journal of Obstetrics
and Gynecology 157(7): 1247-1251.
Kuhnert, B. and Kuhnert, P. 1988. Lead and cadmium concentrations in mother and fetus (letter).
American Journal of Obstetrics and Gynecology 158(1): 220.
Kupfer, D. and Bulger, W.H. 1987. Metabolic activation of pesticides with proestrogenic activity.
Federation Proceedings 46(5): 1864-1869.
Kurita, H., Ludwig, J.P., and Ludwig, M. 1987. Results of the 1987 Michigan colonial waterbird
monitoring project on Caspian terns and double-crested cormorants: egg incubation and
field studies of colony productivity, embryologic mortality, and deformities. Ecological
Research Services, Inc.
Ku'rzel, R.B. and Cetrulo, C.L. 1981. The effect of environmental pollutants on human
reproduction, including birth defects. Environmental Science and Technology 15: 626-
640.
Kutscher, C.L., Sembrat, M., Kutscher, C.S., and Kutscher, N.L. 1985. Effects of the high
methylmercury dose used in the collaborative behavioral teratology study on brain
anatomy. Neurobehavioral Toxicology and Teratology 7(6): 775-777.
Lancaster, J. 1990. Dolphin deaths in Gulf Coast prompt scientific probe. Los Angeles Times.
June 10.
Landers, J.P. and Bunce, N.J. 1991. The Ah receptor and the mechanism of dioxin toxicity.
Biochemical Journal 275: 273-287.
Landrum, P.P., Nihart, S.R., Eadie, B.J., and Herche, L.R. 1987. Reduction in bioavailability of
organic contaminants to the Amphipod, Pontoporeia Hoyi by dissolved organic matter of
sediment interstitial waters. Environmental Toxicology and Chemistry 6: 11-20.
Langworth, S., Elinder, C., and Akesson, A. 1988. Mercury exposure from dental fillings.
Sweden Dental Journal 12: 69-70.
Lapointe, B.E., Littler, M.M., and Littler, D.S. 1987. A comparison of nutrient-limited
productivity in macroalgae from a Caribbean barrier reef and from a mangrove ecosystem.
Aquatic Botany 28: 243-255.
161
-------
Larkfors, L,, Sundberg, J., and Ebendal, T. 1991. Methylmercury induced alterations in the nerve
growth factor level in the developing brain. Brain Research (Developmental Brain
Research) 62:287-291.
L'Arrivee, L. and Blokpoel, H. 1988. Seasonal distribution and site fidelity in Great Lakes
Caspian terns. Colonial Waterbirds 11: 204-214.
Larsson, U.R., Elmgren, R., and Wulff, F. 1985. Eutrophication and the Baltic Sea: Causes and
consequences. Ambio 14: 10-14.
Lauwerys, R., Buchet, J., Roels, H., and Hubermont, G. 1978. Placenta! transfer of lead, mercury,
cadmium, and carbon monoxide in women. Environmental Research 15: 278-289.
Lavigne, D.M. and Schmitz, OJ. 1990. Global wanning and increasing population densities: a
prescription for seal plagues. Marine Pollution Bulletin 21(6): 280-284.
Lean, D.R.S. 1987. Studies on the nutrient status of Lake Ontario. Canadian Journal of Fisheries
and Aquatic Science 44: 2039-2241.
Leatherland, J.F. 1992. Endocrine and reproductive function in Great Lakes Salmon. In: Colborn,
T. and Clement, C. (eds.). Chemically-induced Alterations in Sexual and Functional
Development: The Human-Wildlife Connection. Princeton, NJ: Princeton Scientific
Publishing. In press.
Leatherland, J.F., Donaldson, E.M., Down, N.E., Flett, P.A., Moccia, R., Munkittrick, K.R.,
Sonstegard, RA., and Van der Kraak, G. 1991. Field observations on reproductive and
developmental dysfunction and native salmonids from the Great Lakes. Pp 17-18 in:
Schneider, S. and Campbell, R. (eds.). Cause-Effect Linkages II Symposium Abstracts.
Michigan Audubon Society, Traverse City, MI. September 27-28, 1991.
Leatherland, J.F. and Sonstegard, R. 1982. Thyroid responses in rats fed diets formulated with
Great Lakes Coho salmon. Bulletin of Environmental Contamination and Toxicology 29:
341-346.
Lech J.J., Vodicinik M.J., and Elcombe C.R. 1982. Induction of monooxygenase activity in fish.
Pp. 107-148 in: Weber, L.J. (ed.). Aquatic Toxicology. New York, NY: Raven Press.
Lee, I.D. and Dixon, R.L. 1975. Effects of mercury on spermatogenesis studied by velocity
sedimentation, cell separation and serial mating. Journal of Pharmacology and
Experimental Therapy 194: 171-181.
Lee Y.-Z., Leighton, F.A., Peakall, D.B., Norstrom, RJ., O'Brien, P.J., Payne, J.F., and
Rahimtula, A.D. 1985. Effects of ingestion of Hiberaia and Prudhoe Bay crude oils on
162
-------
hepatic and renal mixed function oxidase in nestling herring gulls (Larus argentatus).
Environmental Research 36: 248-255.
Lee Y.-Z., O'Brien, P.J., Payne, J.F., and Rahimrula, A.D. 1986. Toxicity of petroleum crude oils
and their effect on xenobiotic metabolizing enzyme activities in the chicken embryo in
avo. Environmental Research 39: 153-164.
Lein, A.Yu. and M.V. Ivanov. 1992. Interaction of carbon, sulphur, and oxygen cycles in
continental and marginal seas. In: Howarth, R.W., Stewart, J.W.B., and Ivanov, M.V.
(eds.). Sulphur Cycling on the Continents: Wetlands, Terrestrial Ecosystems, and
Associated Water Bodies. Chichester, United Kingdom: Wiley & Sons, Inc.
Lentnek, M., Griffith, O.W., and Rifkind, A.B. 1991. 2,3,7,8-Tetrachlorodibenzo-p-dioxin
increases reliance on fats as a fuel source independently of diet: evidence that diminished
carbohydrate supply contributes to dioxin lethality. Biochemical and Biophysical Research
Communications 174(3): 1267-1271.
Leoni, V., Fabiani, L., Marinelli, G., Puccetti, G., Tarsitani, G.F., de Carolis, A., Vexcia, N.,
Morini, A., Aleandri, V., Pozzi, V., Cappa, F., and Barbati, D. 1989. PCB and other
organochlorine compounds in blood of women with or without miscarriage: a hypothesis
of correlation. Ecotoxicology and Environmental Safety 17: 1-11.
LeVay, S. 1991. A difference in hypothalamic structure between heterosexual and homosexual
men. Science 253: 1034-1037.
Levin, W., Welch, R.M., and Conney, A.H. 1968. Estrogenic action of DDT and its analogs.
Federation Proceedings 27: 649 (abst 2440).
Li, K. 1988. Lead values in umbilical cord blood and maternal blood. Journal of the Royal
Society of Health 108: 59.
Likens, G.E. 1972. Nutrients and eutrophication. American Society of Limnology &
Oceanography Special Symposium I.
Lin, F.H., Clark, G., Birnbaum, L.S., Lucier, G.W., and Goldstein, J.A. 1991a. Influence of the
Ah locus on the effects on 2,3,7,8-tetrachlorodibenzo-p-dioxin on the hepatic epidermal
growth factor receptor. Molecular Pharmacology 39(3): 307-313.
Lin, F.H., Stohs, S.J., Birnbaum. L.S., Clark,, G., Lucier, G.W., and Goldstein, J.A. 1991b. The
effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the hepatic estrogen and
glucocorticoid receptors in congenic strains of Ah responsive and Ah nonresponsive
C57BL/6J mice. Toxicology and Applied Pharmacology 108(1): 129-139.
163
-------
Lindahl, G. and Wallstrora, K. 1985. Nitrogen fixation (acetylene reduction) in planktonic
cyanobacteria in Oregrundsgrepen, SW Bothnian Sea. Archiv fur Hydrobiologie 104:193-
204.
Linden, J., Pohjanviita, R., Rahko, T., and Tuomisto, J. 1991. TCDD decreases rapidly and
persistently serum melatonin concentration without morphologically affecting the pineal
gland in TCDD-resistant Han/Wistar rats. Pharmacology and Toxicology 69(6): 427-432.
Under, R.E., Gaines, T.B., and Kimbrough, G.D. 1974. The effect of polychlorinated biphenyls
on rat reproduction. Food and Cosmetics Toxicology 12: 67-77.
JJndstrom, H., Luthman, J., Oskarsson, A., Sundberg, J., and Olson, L. 1991. Effects of long-
term treatment with methyl mercury on the developing rat brain. Environmental Research
56: 159-169.
Littler, M.M. and Murray, S.N. 1975. Impact of sewage on the distribution, abundance and
community structure of rocky intertidal macro-organisms. Marine Biology 30: 277-291.
Littler, M.M. and Murray, S.N. 1978. Influence of domestic wastes on energetic pathways in
rocky intertidal communities. Journal of Applied Ecology 15: 583-596.
Littler, M.M., Littler, D.S., and Lapointe, B.E. 1988. A comparison of nutrient- and light-limited
photosynthesis in psammophytic versus epilithic forms of Halimeda (Caulerpales,
Halimedaceae) from the Bahamas. Coral Reefs 6: 219-225.
Lombet, A., Mourre, G, and Lazdunski, M. 1988. Interaction of insecticides of the pyrethroid
family with specific binding sites on the voltage-dependent sodium channel from
mammalian brain. Brain Research 459(1): 44-53.
Lommel, A., Kruse, H., and Wasserman, O. 1985. Organochlorines and mercury in blood of a
fish-eating population at the River Elbe in Schleswig-Holstein, FRG. Archives of
Toxicological Supplements 8: 264-268.
van Loveren, H., Krajnc, E.I., Rombout, P.J.A., Blommaert, F.A., and Vos, J.G. 1990. Effects
of ozone, hexachlorobenzene, and Bis(tri-n-butyltin) oxide on natural killer activity in
the rat lung. Toxicological Applications Pharmacology 102: 21-33.
Ludwig, J.P. and Tomoff, C. 1966. Reproductive success and insecticide residues in Lake
Michigan herring gulls. Jack-Pine Warbler 44(2): 77-84.
Ludwig, J.P. 1984. Decline, resurgence, and population dynamics of Michigan and Great Lakes
double-crested cormorants. Jack-Pine Warbler 62(4): 91-102.
164
-------
Ludwig, J.P. and Giesy, J.P. 1990. Effects of PCBs and other halogenated aromatic hydrocarbons
on Caspian tern reproduction in the Upper Great Lakes. A research proposal.
Unpublished.
Ludwig, J.P. 1992. Senior Ecologist and President, Ecological Research Services (ERS), Bay
City, ML
Lundberg, C. 1973. Effects of long-term exposure to DDT on the oestrus cycle and the
frequency of implanted ova in the mouse. Environmental Physiology and Biochemistry
3: 127-131.
Lundberg, K., Gronvik, K.O., and Dencker, L. 1991. 2,3,7,8-Tetrachlorodibenzo-p-dioxin
(TCDD) induced suppression of the local immune response. International Journal of
Immunopharmacology 13(4): 357-368.
Lundholm, E. 1987. Thinning of eggshells in birds by DDE: mode of action on the eggshell
gland. Comparative Biochemistry and Physiology 88C(1): 1-22.
Lundkvist, U. and Kindahl, H. 1989. Plasma concentrations of 15-keto-13, 14-dihydro-PGF-2
alpha, oestrone sulphate, oestradiol-17 beta and progesterone in pregnant guinea-pigs
treated with polychlorinated biphenyls. Journal of Reproduction and Fertility 87(1): 55-
62.
Lustick, S., Voss, T., and Peterle, T. 1973. Effects of DDT on steroid metabolism and energetics
in bobwhite quail (Colinus virginianus). Pp. 213-233 in: Morrison, J.A. and Lewis, J.C.
(eds.). First National Bobwhite Quail Symposium. Stillwater, Oklahoma, OK: University
Press.
Mably, T.A., Moore, R.W., Goy, R.W., and Peterson, R.E. 1992. In utero and lactational
exposure of male rats to 2,3,7,8-tetrachIorodibenzo-p-dioxin. 2. Effects on sexual
behavior and the regulation of luteinizing hormone secretion in adulthood. Toxicology and
Applied Pharmacology 114: 108-117.
Madge, D.S. 1977. Effects of trichlorophenoxyacetic acid and chlorodioxins on small intestinal
function. General Pharmacology 8: 319-324.
Mahanty, H.K., Fineran, BA., and Gresshoff, P.M. 1983. Effects of polychlorinated biphenyls
(Aroclor 1242) on the ultrastmcture of certain planktonic algae. Botanical Gazette 144(1):
56-61.
Malone, T.C. 1982. Factors influencing the fate of sewage-derived nutrients in the lower Hudson
estuary and New York Bight. In: Mayer, G.F. (ed.). Ecological Stress and the New York
Bight: Science and Management. Columbia, SC: Estuarine Research Federation.
165
-------
Manchester, D., Gordon, S., Golas, C, Roberts, E., and Okey, A. 1987. Ah receptor in human
placenta: stabilization by molybdate and characterization of binding of 2,3,7,8-
tetrachlorodibenzo-p-dioxin,3-methylcholanthrene,andbenzo(a)pyrene.CancerResearch
47(18): 4861-4868.
Manis, J. and Kim, G. 1979. Introduction of iron transport by a potent inducer of aryl
hydrocarbon hydroxylase, 2,3,7,8-tetrachlorodibenzo-p-dioxin. Archives of
Environmental Health 34(3): 141-145.
Manz, A., Berger, J., Dwyer, J.H., Flesch-Janys, D., Nagel, S., and Waltsgott, H. 1991. Cancer
mortality among workers in chemical plant contaminated with dioxin. Lancet 338,8873:
959-964.
Marino, R., Howarth, R.W., Shamess, J., and Prepas, E.E. 1990. Molybdenum and sulfate as
controls on the abundance of nitrogen-fixing cyanobacteria in saline lakes in Alberta.
Limnology & Oceanography 35: 245-259.
Marks, G.S. 1985. Exposure to toxic agents: the heme biosynthetic pathway and hemoproteins
as indicators. CRC Critical Review of Toxicology 15: 151-179.
Martin, S.G., Thiel, DA., Duncan, J.W., and Lance, W.R. 1987. Effects of a paper industry
sludge containing dioxin on wildlife in red pine plantations. Pp. 363-377. Technical
Association of Pulp and Paper Industries (TAPPI) Proceedings. 1987 Environmental
Conference. Portland, OR.
Martineau, D., Lagace, A., Beland, P., Higgins, R., Armstrong, D., and Shugart, L.R. 1988.
Pathology of stranded beluga whales (Delphinapterus leucas) from the St. Lawrence
estuary, Quebec, Canada. Journal of Comparative Physiology 98: 287-311.
Martineau, D., Beland, P., Desjardins, C., and Lagace, A. 1987. Levels of organochlorine
chemicals in tissues of beluga whales (Delphinapterus leucas) from the St. Lawrence
estuary, Quebec, Canada. Archives of Environmental Contamination and Toxicology 16:
137-147.
Martineau, D., Beland, P., Desjardins, C., and Vezina, A. 1985. Pathology, toxicology, and
effects of contaminants on the population of the St. Lawrence beluga (Delphinaterus
leucas). Quebec, Canada. ICES: CM. 1985.
Martinez, E.M. and Swartz, W J. 1992. Effects of methdxychlor on the reproductive system of
the adult female mouse: II. Ultrastructural observations. Reproductive Toxicology 6(1):
93-98.
166
-------
Mason, G., Sawyer, T, Keys, B., Bandiera, S., Romkes, M., Piskorska-Pliszczynska, J.,
Smudzka, B., and Safe, S. 1985. Polychlorinated dibenzofurans (PCDFs): correlation
between in vivo and in vitro structure-activity relationships. Toxicology 37: 1-12.
Mason, G., Farrell, K., Keys, B., Piskorska-Pliszczynska, J., Safe, L., and Safe, S. 1986.
Polychlorinated dibenzo-p-dioxins: quantitative in vitro and in vivo structure activity
relationships. Toxicology 41: 21-31.
Mason, G., Zacharewski, T., Denomme, M., Safe, L., and Safe, S. 1987. Polybrominated
dibenzo-p-dioxins and related compounds: quantitative in vivo and in vitro structure
activity relationships. Toxicology 44: 245-255.
Mason, R.R. and Schulte, G.L. 1980. Estrogen-like effects of o,p'DDT on the progesterone
receptor of rat uterine cytosol. Research Communications hi Chemical Pathology and
Pharmacology 29: 281-290.
Masuda, Y., Kagawa, R., Kuroki, H., Kuratsune, M., Yoshimura, T., Taki, I., Kusuda, M.,
Yamashita, F., and Hayashi, M. 1978. Transfer of polychlorinated biphenyls from mothers
to fetuses and infants. Bulletin of Environmental Contamination and Toxicology 16: 543-
546.
May, E.B., Lukacovic, R., King, H., and Lipsky, M.M. 1987. Hyperplastic and neoplastic
alterations in the livers of white perch (Morone americana) from the Chesapeake Bay.
Journal of the National Cancer Institute 79: 137-143.
McArthur, M.L.B., Fox, GA., Peakall, D.B., and Philogene, BJ.R. 1983. Ecological significance
of behavioral and hormonal abnormalities in breeding ring doves fed an organochlorine
chemical mixture. Archives of Environmental Contamination and Toxicology 12: 343-
353.
McComb, A.J., Atkins, R.P., Birch, P.B., Gordon, D.M., and Luketelich, R.J. 1981.
Eutrophication in the Peel-Harvey estuarine system, Western Australia. In: Nielson, B J.
and Cronin, L.E. (eds.) Estuaries and Nutrients. Humana, NY.
McConkey, D.J., and Orrenius, S. 1989. 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) kills
glucocorticoid-sensitive thymocytes in vivo. Biochemistry and Biophysical Research
Communications 160(3): 1003-1008.
McCormack, K.M., Arneric, S.P., and Hook, J.B. 1979. Action of exogenously administered
steroid hormones following perinatal exposure to polybrominated biphenyls. Journal of
Toxicology and Environmental Health 5: 1085-1094.
167
-------
McGlathery, KJ., Howarth, R.W., and Marino, R. 1992. Nutrient limitation of the macroalga,
Penicillus capitatus, associated with subtropical seagrass meadows in Bermuda. Estuaries
15: 18-25. In press.
McLachlan, JA. 1985. Estrogens in the Environment, n. Influences on development. New York.
Elsevier Science Publishing Company.
McNulty, W.P. 1984. Fetotoxicity of 2,3,7,8-tetracMorodibenzo-p-dioxin (TCDD) for rhesus
macaques (Macaco mulatto). American Journal of Primatology 6: 41-47.
Mearns, AJ., Raines, E., Klepple, G.S., McGrath, RA, McLaughlin, J.JA., Segar, DA., Sharp,
J.H., Walsh, JJ., Word, J.Q., Young, D.K., and Young, M.W. 1982. Effects of nutrients
and carbon loadings on communities and ecosystems. In: Mayer, G.F. (ed.). Ecological
Stress and the New York Bight: Science and Management. Columbia, SC. Esruarine
Research Federation.
Menconi, S., Clark, J.M., Langenbert, P., and Hryhorczuk, D. 1988. A preliminary study of
potential human health effects in private residences following chlordane application for
termite control. Archives of Environmental Health 43(5): 349-352.
Mes, J. and Davies, D. 1979. Presence of polychlorinated biphenyl and organochlorine pesticide
residues and the absence of polychlorinated terphenyls in Canadian human milk samples.
Bulletin of Environmental Contamination and Toxicology 21: 381-387.
Mes, J., Doyle, J., Adams, B., Davies, D., and Turton, D. 1984. Polychlorinated biphenyls and
organochlorine pesticides in milk and blood of Canadian women during lactation. Archives
of Environmental Contamination and Toxicology 13: 217-223.
Mes, J., Davies, D., Turton, D., and Sun, W. 1986. Levels and trends of chlorinated hydrocarbon
contaminants in the breast milk of Canadian women. Food Additives and Contaminants
3: 313-322.
Mes, J., Turton, D., Davies, D., Sun, W., Lau, P., and Weber, D. 1987. The routine analysis of
some specific isomers of polychlorinated biphenyl congeners in human milk. International
Journal of Environmental Analytical Chemistry 28: 197-205.
Miller, D.S., Peakall, D.B., and Kinter, W.B. 1978. Ingestion of crude oil: sublethal effects in
herring gull chicks. Science 199: 315-317.
Mineau, P. and Weseloh, D. 1981. Low-disturbance monitoring of herring gull reproductive
success on the Great Lakes. Colonial Waterbirds 4: 138-142.
Mineau, P., Fox, G., Norstrom, R., Weseloh, D., Hallett, D., and Ellenton, J. 1984. Using the
herring gull to monitor levels and effects of organochlorine contamination in the Canadian
168
-------
Great Lakes. Pp. 426-452 in: Nriagu, J. and Simmons, M. (eds.). Toxic Contaminants in
the Great Lakes. John Wiley & Sons.
Miura, K. and Imura, N. 1987. Mechanism of methylmercury cytotoxicity. Critical Reviews in
Toxicology 18: 161-188.
Moccia, R., Fox, G., and Britton, A. 1986. A quantitative assessment of thyroid histopathology
of herring gulls (Larus argentatus) from the Great Lakes and a hypothesis on the causal
role of environmental contaminants. Journal of Wildlife Disease 22: 60-70.
Moccia, R.D., Leatherland, J.F., and Sonstegard, R.A. 1981. Quantitative interlake comparison
of thyroid pathology in Great Lakes Coho (Onchorhynchus Jdsutch) and chinook
(Onchorhynchus tschawytschd) salmon. Cancer Research 41: 2200-2210.
Mohammed, A., Halberg, E., Rydstrom, J., and Slanina, P. 1985. Toxaphene: accumulation in the
adrenal cortex and effect on ACTH-stimulated corticosteroid synthesis in the rat.
Toxicology Letters. 24(2-3): 137-143.
Molot, L.A. and Dillon, P.J. 1991. Nitrogen/phosphorus ratios and the prediction of chlorophyll
in phosphorus-limited lakes in central Ontario. Canadian Journal of Fisheries and Aquatic
Science 48: 140-145.
Moore, S.A., Jr. and Harris, R.C. 1972. Effects of polychlorinated biphenyl on marine
phytoplankton communities. Nature 240: 356-357. December 8.
Moore, R.W., Potter, C.L., Theobald, H.M., Robinson, J.A., and Peterson, R.E. 1985.
Androgenic deficiency in male rats treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin.
Toxicology and Applied Pharmacology 79: 99-111.
Moore, R.W. and Peterson, R.E. 1988. Androgen catabolism and excretion in 2,3,7,8-
tetrachlorodibenzo-p-dioxin-treated rats. Biochemical Pharmacology 37(3): 560-562.
Moore, R.W., Parsons, J.A., Bookstaff, R.C., and Peterson, R.E. 1989. Plasma concentrations of
pituitary hormone in 2,3,7,8-tetrachlorodibenzo-p-dioxin-treated male rats. Journal of
Biochemistry and Toxicology 4(3): 165-172.
Moorhead, D.L and Kosinski, RJ. 1986. Effect of Atrazine on the productivity of artificial
stream algal communities. Bulletin of Environmental Contamination and Toxicology 37:
330-336.
Morin, A., Hambright, K.D., Hairston, N., Sherman, D., and Howarth, R.W. 1991. Consumer
control of gross primary production in replicate freshwater ponds. Verhandlunge der
Intemationalen Vereinigung fur Theoretische und Angewandte Limnologie. In press.
169
-------
Morris, D.L., Jordan S.D., and Holsapple, M.P. 1991. Effects of 2,3,7,8-terrachlorodibenzo-p-
dioxin (TCDD) on humoral immunity: I. Similarities to Staphylococcus aureus I (SAC)
in the in vitro T-dependent antibody response. Immunopharmacology 21(3): 159-169.
Morse, J.W., Zullig, JJ., Bernstein, L.D., Millero, FJ., Milne, P., Mucci, A., and Choppin, G.R.
1985. Chemistry of calcium carbonate-rich shallow water sediments hi the Bahamas.
American Journal of Science 285: 147-185.
Moser, G J. and Smart, R.C. 1989. Hepatic tumor-promoting chlorinated hydrocarbons stimulate
protein kinase C activity. Carcinogenesis 10(5): 851-856.
Mosser, J.L., Fisher, N.S., and Wurster, C.F. 1972. Polychlorinated biphenyls and DDT alter
species composition hi mixed cultures of algae. Science 176: 533-535. May 5.
Muir, D.C.G., Ford, CA., Stewart, R.EA., Smith, T.G., Addison, R.F., Zinck, M.E., and Beland,
P. 1990. Organochlorine contaminants in belugas (Delphinapterus leucas) from Canadian
waters. Canadian Bulletin of Fisheries and Aquatic Science 224: 165-190.
Mukhtar, H., Kumar, A., Husain, M.M., and Krishna Murti, C.R. 1981. Aryl hydrocarbon
hydroxylase in pigeon skin and its possible relevance to monitoring air pollution.
Ecotoxicology and Environmental Safety 5: 97-105.
Murdoch, P.S. and Stoddard, J.L. 1991. The role of nitrate hi the acidification of streams in the
Catskiil Mountains of New York. Report to EPA.
Murphy, T.J. and Rzeszutko, C.P. 1977. Precipitation inputs of PCBs to Lake Michigan. Journal
of Great Lakes Research 3: 305-312.
Murphy, R.S., Kutz, F.W., and Strassman, S.C. 1983. Selected pesticide residues or metals in
blood or urine specimens from a general population survey. Environmental Health
Perspectives 48: 81-86.
Murphy, T.J. 1984. Atmospheric inputs of chlorinated hydrocarbons to the Great Lakes. Pp. 54-
79 hi: Nriagu, J.O. and Simmons, M.S. (eds.). Toxic Contaminants in the Great Lakes.
New York, NY: John Wiley & Sons.
Murphy T.J., Paolucci, G., Schinsky, A., Combs, M., and Pokojowczyk, J. 1982. Inputs of PCB
from the atmosphere to Lakes Huron and Michigan. Report of U.S. EPA Project R-
805325. Duluth Environmental Research Laboratory. Cited hi Murphy (1984).
Murphy, T.J. and Schinsky, A.L. 1983. Net atmospheric inputs of PCBs to the ice cover of Lake
Huron. Journal of Great Lakes Research 9: 92-96.
170
-------
Murphy, R.S., Kutz, F.W., and Strassman, S.C. 1983. Selected pesticide residues or metals in
blood or urine specimens from a general population survey. Environmental Health
Perspectives 48: 81-86.
Murray, F.J., Smith, F.A., Nitschke, K.D., Humison, CO., Kociba, RJ., and Schwetz, BA. 1979.
Three-generation reproduction study of rats given 2,3»7,8-tetrachlorodibcnzo-p-dioxin
(TCDD) in the diet. Toxicology and Applied Pharmacology 50: 241-252.
Muzi, G., Gorski, J.R., and Rozman, K. 1987. Composition of diet modifies toxicity of 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD) in cold-adapted rats. Archives of Toxicology,
61(1): 34-39.
Myers, V.B. and Iverson, R.I. 1981. Phosphorus and nitrogen limited phytoplankton productivity
in northeastern Gulf of Mexico coastal estuaries. In: Nielson, B J. and Cronin, L.E. (eds.).
Estuaries and Nutrients. Humana, NY.
Mykkanen, H., Rasanen, M., and Kimppa, S. 1986. Dietary intakes of mercury, lead, cadmium
and arsenic by Finnish children. Human Nutrition: Applied Nutrition 40A: 32-39.
Nagaoka, S., Kamuro, H., Oda, H., and Yoshida, A. 1991. Effects of polychlorinated biphenyls
on cholesterol and ascorbic acid metabolism in primary cultured rat hepatocytes.
Biochemical Pharmacology 41(8): 1259-1261.
Narasimhan, T.R., Safe, S., Williams, H.J., and Scott A.I. 1991. Effects of 2,3,7,8-
tetrachlorodibenzo-p-dioxin on 17 beta-estradiol-induced metabolism in MCF-7 human
breast cancer cells: 13C nuclear magnetic resonance spectroscopy studies. Molecular
Pharmacology 40(6): 1029-1035.
Narbonne, J.F., Garrigues, P., Ribera, D., Raoux, C, Mathieu, A., Lemaire, P., Salaun, J.P., and
Lafaurie, M. 1991. Mixed-function oxygenase enzymes as tools for pollution monitoring:
field studies on the French coast of the Mediterranean sea. Comparative Biochemistry and
Physiology 100C: 37-42.
National Oceanic and Atmospheric Administration (NOAA) USEPA. 1988. Strategic assessment
of near coastal waters: northeast case study. Susceptibility and status of northeast estuaries
to nutrient discharges. Rockville, MD.
National Research Council (NRQ. 1993. Report of the Committee on Wastewater Management
for Coastal Urban Areas, Water, Science, and Technology Board. Washington, DC. In
review.
National Oceanic and Atmospheric Association (NOAA). 1991. Environmental Conservation
Division Briefing Book: Programs and Accomplishments 1986-1991. Environmental
Conservation Division.
171
-------
Nebert, D.E., Eisen, HJ., Negishi, M., Lang, MA., and Hjelmeland, L.M. 1981. Genetic
mechanisms controlling the induction of polysubtrate monooxygenase (P-450) activities.
Annual Review of Pharmacology and Toxicology 21: 431-462.
Nebert, D.W. and Gonzalez, F.J. 1987. P450 genes: structure, evolution and regulation. Annual
Review of Biochemistry 56: 945-993.
Nellbring, S., Hansson, S., Aneer, G., Westin, L. 1980. Impact of oil on local fish fauna. In: The
Tsesis Oil Spill. Kineman, J.J., Elmgren, R., and Hanson, S. (eds.). U.S. Department of
Commerce. NOAA.
Nelson, J.A. 1974. Effects of dichlorodiphenyltrichloroethane (DDT) analogs and polychlorinated
biphenyl (PCS) mixtures on 17B-[3H] estradiol binding to rat uterine receptor.
Biochemical Pharmacology 23: 447-451.
Nelson, J.A., Stuck, R.F., and James, R. 1976. Estrogenically active forms of o,p'DDT and
methoxychlor. Pharmacologist 18: 247. (Abst. 730).
Nelson, J.A., Stuck, R.F., and James, R. 1978. Estrogenic activities of chlorinated hydrocarbons.
Journal of Toxicology and Environmental Health. 4: 325-340.
Nelson, L. 1990. Pesticide perturbation of sperm cell function. Bulletin of Environmental
Contamination and Toxicology 45: 876-882.
Neubert, D. and Dillman, I. 1972. Embryotoxic effects in mice treated with 2,4,5-
trichlorophenoxy acetic acid and 2,3,7,8-tetrachlorodibenzo-p-dioxin. Nauyn-
Schmiedeberg's Archives of Pharmacology 272: 243-264.
Neubert, R., Jacob-Muller, U., Stahlmann, R., Helge, H., and Neubert, D. 1990.
Polyhalogenated dibenzo-p-dioxins and dibenzofurans and the immune system. I. Effects
on peripheral lymphocyte subpopulations of a non-human primate (Callithrix jacchus)
after treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Archives of
Toxicology 64(5): 345-359.
Nikolaidis, E.B., Brunstrom, B., and Denker, L. 1988. Effects of TCDD congeners 3,3'4,4'-
tetrachlorobiphenyl and 3,3'4,4'-tetrachloroazoxybenzene on lymphoid development in the
Bursa of Fabricius in the chick embryo. Toxicology and Applied Pharmacology 92: 315-
323.
Nisbet, I.C.T. and Paxton, M.B. 1982. Statistical aspects of three-generation studies of the
reproductive toxicity of TCDD and 2,4,5,-T. American Statistician 36(3): 290-298.
Nisbet, I.C.T. and Drury, W.H. 1984. Super-normal clutches in herring gulls in New England.
The Condor 86: 87-89.
172
-------
Nixon, S.W., Kelly, J.R., Furnas, B.N., Oviatt, C.A., and Hale, S.S. 1980. Phosphorus
regeneration and the metabolism of coastal marine bottom communities. In: Tenore, K.R.
and Coull, B.C. (eds.). Marine Benthic Dynamics. Columbia, SC: University of South
Carolina Press.
Nixon, S.W., Oviatt, C, Frithsen, J., and Sullivan, B. 1986. Nutrients and productivity of
estuaries and coastal marine ecosystems. Journal of the Limnology Society of South
Africa 12: 43-71.
Nixon, S.W. 1988. Physical energy inputs and the comparative ecology of lake and marine
ecosystems. Limnology & Oceanography 33: 1005-1025.
Nixon, S.W. 1992. Quantifying the relationship between nitrogen input and the productivity of
marine ecosystems. Advances in Marine Technology Conference 5: 57-83.
Nordberg, G. 1988. Current concepts in the assessment of effects of metals in chronic low-level
exposures-considerations of experimental and epidemiological evidence. The Science Of
the Total Environment 71: 243-252.
Noren, K. 1983. Levels of organochlorine contaminants in human milk in relation to the dietary
habits of the mothers. Acta Paediatric Scandinavia 72(6): 811-816.
Norin, L.L. 1977. 14C-bioassays with the natural phytoplankton in the Stockholm archipelago.
Ambio Special Report 5: 15-21.
Norstrom, RJ., Hallett, D.J., Onuska, F.I., and Comba, M.E. 1980. Mirex and its degradation
products in Great Lakes herring gulls. Environmental Science and Technology 14: 860-
866.
Nosek, J.A., Craven, S.R., Sullivan, J.R., Olson, J.R., and Peterson, R.E. 1992. Metabolism and
disposition of 2,3,7,8-tetrachlorodibenzo-p-dioxin in ring-necked pheasant hens, chicks,
and eggs. Journal of Toxicology and Environmental Health 35(3): 153-164.
O'Connors, H.B., Jr., Wurster, C.F., Powers, C.D., Biggs, D.C., and Rowland, R.G. 1978.
Polychlorinated biphenyls may alter marine trophic pathways by reducing phytoplankton
size and production. Science 201: 737-739. August 25.
Oehme, M., Ryg, M., Furst, P., Furst, C, Meemken, HA., and Groebel, W. 1990. Re-evaluation
of concentration levels of polychlorinated dibenzo-p-dioxins and dibenzofurans in Arctic
seals from Spitzenbergen. Chemosphere 21(4-5): 519-523.
Officer, C.B. and Ryther, J.H. 1980. The possible importance of silicon in marine eutrophication.
Marine Ecology Progress Series 3: 83-91.
173
-------
Officer, C.B., Biggs, R.B., Taft, J., Cronin, L.E., Tyler, M.A., and Boynton, W.R. 1984.
Chesapeake Bay anoxia: origin, development, and significance. Science 223: 22-27.
O'Kusky, J.R., Boyes, B.E., and McGeer, E.G. 1988. Methylmercury-induced movement and
postural disorders in developing rat: regional analysis of brain catecholamines and
indoleamines. Brain Research 439(1/2): 138-146.
Olie, K., van den Berg, M., and Hutzinger, O. 1983. Formation and fate of PCDD and PCDF
from combustion processes. Chemosphere 12: 627.
Olie, K., Vermeulen, P., and Hutzinger, O. 1977. Chlorodibenzo-p-dioxins and
chlorodibenzofurans are trace components of fly ash and flue gas of some municipal
incinerations in The Netherlands. Chemosphere 6: 455.
van der Oost, R., Heida, K., Opperhuizen, A., and Vermeulen, N.P.E. 1991. Interrelationships
between bioaccumulation of organic trace pollutants (PCBs, organochlorine pesticides and
PAHs), and MFO-induction in fish. Comparative Biochemistry and Physiology 100C: 43-
47.
Orberg, J. and Kihlstroem, J.E. 1973. Effects of long-term feeding of polychlorinated biphenyls
(PCB, Qopen A 60) on the length of the oestrous cycle and on the frequency of
implanted ova in the mouse. Environmental Research 6: 176-179.
Ousterhout, J.M., Struck, R.F., and Nelson, J.A. 1979. Estrogenic properties of methoxychlor
metabolites. Federation Proceedings 38: 537. (Abst).
Ousterhout, J.M., Struck, R.F., and Nelson, J.A. 1981. Estrogenic activities of methoxychlor
metabolites. Biochemical Pharmacology 30: 2868-258.
Oviatt, CA., Keller, A., Sampou, P.A., and Beatty, L.L. 1986. Patterns of productivity during
eutrophication: a mesocosm experiment. Marine Ecology Progress Series 28: 69-80.
Office of Water Regulations and Standards (OWRS). Work/Quality Assurance Project Plan for
the Bioaccumulation Study. U.S. Environmental Protection Agency, July 1986.
Paerl, H.W., Crocker, K.M., and Prufert, L.E. 1987. Limitation of N2 fixation hi coastal marine
waters: relative importance of molybdenum, iron, phosphorus, and organic matter
availability. Limnology & Oceanography 32: 525-536.
Paerl^H.W. and Carlton, R.C. 1988. Control of nitrogen fixation by oxygen depletion in surface-.
associated microzones. Nature 332: 260-262.
Parker, C.A. and O'Reilly, J.E. 1991. Oxygen depletion in Long Island Sound: a historical
perspective. Estuaries 14: 248-264.
174
-------
Parsons, A.H. and Peterle, T.J. 1977. DDE and avian eggshell thinning: ultrastructural evidence
of decreased parathyroid activity. Poultry Science 56: 1745.
Pastorak, R.A. and Bilyard, G.R. 1985. Effects of sewage pollution on coral-reef communities.
Marine Ecology Progress Series 21: 175-189.
Payne, J.F., Fancey, L.L., Rahimtula, A.D., and Porter, E.L. 1987. Review and perspective on the
use of mixed-function oxygenase enzymes in biological monitoring. Comparative
Biochemistry and Physiology 86C: 233-235.
Peakall, D.B. 1967. Pesticide-induced enzyme breakdown of steroids in birds. Nature 216: 505-
506.
Peakall, D.B. 1970a. Pesticides and the reproduction of birds. Scientific American 222: 72-78.
Peakall, D.B. 1970b. p,p'DDT: effect on calcium metabolism and concentration of estradiol in
the blood. Science 168: 592-594.
Peakall, D.B. 1976. DDT in rainwater in New York following applications in the Pacific
Northwest. Atmospheric Environment 10: 899-900.
Peakall, D.B., Fox, G.A., Oilman, A.P., Hallett, DJ., and Norstrom, RJ. 1980. The herring gull
as a monitor of Great Lakes contamination. Pp. 337-344 in: Afghan, B.K. and Mackay,
D. (eds.). Hydrocarbons and halogenated hydrocarbons in the aquatic environment. New
York, NY: Plenum Press.
Peakall, D.B. and Fox, G.A. 1987. Toxicological investigations of pollutant-related effects in
Great Lakes gulls. Environmental Health Perspective 71: 187-193.
Peakall, D.B. 1988. Known effects of pollutants on fish-eating birds in the Great Lakes of North
America. Pp. 39-54. Proceedings, Chronic Effects of Toxic Contaminants in Large Lakes,
Vol 1. World Conference on Large Lakes, Mackinac Island, MI. May 1986.
Peckham, N.H. and Choi, B.H. 1986. Surface change alterations in mouse fetal astrocytes due
to methylmercury: an ultra-structural study with cationized ferritin. Experimental
Molecular Pathology 44: 230-234.
Peel, DA. 1975. Organochlorine residues in antarctic Snow. Nature 154: 324-325.
Pelletier, L., Rossert. J., Pasquier, R., Vial, M.C., and Druet, P. 1990. Role of CD8+ cells in
mercury-induced antoimmunity or immunosuppression in the rat. Scandinavian Journal
of Immunology 31: 65-74.
175
-------
Pesoncn, M., Goksoyr, A., and Andersson, T. 1992. Expression of P4501A1 in a primary culture
of rainbow trout hepatocytes exposed to beta-naphthoflavone or 2,3,7,8-
tetrachlorodibenzo-p-dioxin. Archives of Biochemistry and Biophysics 292(1): 228-233.
Peterle, TJ. 1969. DDT in antarctic Snow. Nature 224: 620.
Peterle, TJ., Lustick, S.I., Nauman, L.E. 1974. Some physiological effects of dietary DDT on
mallard, bobwhite quail, and domestic rabbits. Transactions of the International Congress
on Game Biology 11: 457-478.
Peterle, T.J. 1991. Wildlife Toxicology. New York, NY: Van Nostrand Reinhold.
Peterson, R.E., Seefeld, M.D., Christian, B J., Potter, C.L., Kelling, C.K., and Keesey, RE. 1984.
Pp. 291-308 in: Poland, A. and Kimbrough, R.D. (eds.). The wasting syndrome in
2,3,7,8-tetrachlorodibenzo-p-dioxin toxicity: Basic features and their interpretation.
Banbury Report 18. Cold Spring Harbor Laboratory.
Peterson, R.E., Moore, R.W., Mably, T.A., Bjerke, D.L., and Goy, R.W. 1992. Male reproductive
system ontogeny: effects of perinatal exposure to 23>7,8-tetrachlorodibenzo-p-dioxin.
In: Colborn, T. and Clement, C. (eds.). Chemically Induced Alterations in Sexual and
Functional Development: The Wildlife/Human Connection. Princeton Scientific Publishing,
Inc. In press.
Pils, C. 1987. The 1986-87 Otter Tagging Report. Wisconsin Department of Natural Resources.
Bureau of Wildlife Management. August.
Pippard, L. 1985. Status of the St. Lawrence River population of beluga (Delphinapterus leucas).
Canadian Field-Naturalist 99(3): 438-450.
»
Pohjanvirta, R., Tuomisto, L., and Tuomisto, J. 1989. The central nervous system may be
involved in TCDD toxicity. Toxicology 58: 167-174.
Pohjanvirta, R., Kulju, T., Morselt, A.F., Tuominen, R., Juvonen, R., Rozman, K., Mannisto, P.,
Collan, Y., Sainio, E.L., and Tuomisto, J. 1989a. Target tissue morphology and serum
biochemistry following 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) exposure in a
TCDD-susceptible and TCDD-resistant rat strain. Fundamental and Applied Toxicology
12(4): 698-712.
Pohjanvirta, R., Tuomisto, L., and Tuomisto, J. 1989b. The central nervous system may be
involved in TCDD toxicity. Toxicology 58(2): 167-174.
Pohjanvirta, R. and Tuomisto, J. 1990a. Remarkable residual alterations in responses to feeding
regulatory challenges in Han/Wistar rats after recovery from the acute toxicity of 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD). Food Chemistry and Toxicology 28(1): 677-686.
176
-------
Pohjanvirta, R. and Tuomisto, J. 1990b. 2,3,7,8-Tetrachlorodibenzo-p-dioxin enhances
responsiveness to post-ingestion satiety signals. Toxicology 63(3): 285-299.
Poland, A. and Knutson, J.C. 1982.2,3,7,8-Tetrachlorodibenzo-p-dioxin and related halogenated
aromatic hydrocarbons: examination of the mechanism of toxicity. Annal Review of
Pharmacology and Toxicology 22: 517-554.
Polishuk, Z.W., Wasserman, D., Wasserman, W., Cucos, S., and Ron, M. 1977. Organochlorine
compounds in mother and fetus during labor. Environmental Research 13: 278-284.
Pusey, C.D., Bowman, C, Morgan, A., Weetman, A.P., Hartley, B., and Lockwood, C.M. 1990.
Kinetics and pathogenicity of autoantibodies induced by mercuric chloride in the brown
Norway rat. Clinical and Experimental Immunology 81: 76-82.
Postupalsky, S. 1971a. Bald eagle and osprey study in Ontario. Correspondence to survey co-
operators. October 25.
Postupalsky, S. 1971b. Toxic chemicals and declining bald eagles and cormorants in Ontario.
Canadian Wildlife Service Manuscript, Report No. 20.
Postupalsky, S. 1976. Toxic chemicals and cormorant populations in the Great Lakes. Paper
presented at the Fish Eating Birds Conference. December 2-3, 1976.
Postupalsky, S. 1980. 1980 bald eagle and osprey nesting surveys in Michigan. Report to
Michigan Department of Natural Resources.
Postupalsky, S. 1983.1983 bald eagle and osprey nesting surveys in Michigan. Wildlife Division
Report No. 2964. December 5, 1983.
Postupalsky, S. 1985. 1985 bald eagle and osprey nesting surveys in Michigan. Report to
Michigan Department of Natural Resources.
Potter, C.W. 1992. Collection Manager for Marine Mammals, National Museum of Natural
History, Smithsonian Institution, Washington, DC.
Powell, G.V.N., Kenworthy, W.J., and Fourqurean, J.F. 1989. Experimental evidence for nutrient
limitation of seagrass growth in a tropical estuary with restricted circulation. Bulletin of
Marine Science 44: 324-340.
Powers, C.D., Rowland, R.G., O'Connors, H.B., Jr., and Wurster, C.F. 1977. Response to
polychlorinated biphenyls of marine phytoplankton isolates cultured under natural
conditions. Applied and Environmental Microbiology 35(6): 760-764.
Prego, R. 1992. Flows and budgets of nutrient salts and organic carbon hi relation to a red tide
in the Ria of-Vigo (NW Spain). Marine Ecology Progress Series 79: 289-302.
177
-------
Price, K.S., Flemer, DA., Taft, J.L., and Mackierman, G.B. 1985. Nutrient enrichment of
Chesapeake Bay and its impact on the habitat of striped bass: a speculative hypothesis.
Transactions of the American Fisheries Society 114: 97-106.
Pryor, G.T., Uyeno, E.T., Tilson, HA., and Mitchell, C.L. 1983. Assessment of chemicals using
a battery of neurobehavioral tests: a comparative study. Neurobehavioral Toxicology and
Teratology 5: 91-117.
Pulliaincn, E., Korhonen, K., Kankaanranta, L., and Maki, K. 1992. Non-spawning burbot on the
northern coast of the Bothnian Bay. Ambio 21(2): 170-175.
Quandt, F.M., Kato, E., and Narahashi, T. 1982. Effects of methylmercury on electrical responses
of neuroblastoma cells. Neurotoxicology 3: 205-220.
Quinn, F. 1992. Hydraulic residence times for the Laurentian Great Lakes. Journal of Great
Lakes Research 18: 22-28.
Raga, J.A. and Aguilar, A. 1991. Mass mortality of striped dolphins in Spanish Mediterranean
waters. Pp. 21-25 in: Pastor, X. and Simmonds, M. (eds.). The Mediterranean Striped
Dolphin Die-Off. Proceedings of the Mediterranean striped dolphin mortality International
Workshop, Palma de Mallorca, 4-5 November, 1991.
Rahel, F.J. 1981. Selection for zinc tolerance in fish: results from laboratory and wild
populations. Transactions of the American Fisheries Society 110: 19-28.
Ratcliffe, DA. 1967. Decrease in eggshell weight in certain birds of prey. Nature 215: 208-210.
Rattner, BA., Eroschenko, V.P., Fox, GA., Fry, D.M., and Gorsline, J. 1984. Avian endocrine
responses to environmental pollutants. Journal of Experimental Zoology 232: 683-689.
Rattner, BA. and Ottinger, MA. 1984. Reduced plasma LH concentration in quail exposed to
the organophosphorus insecticide parathion. Journal of Steroid Biochemistry 20: 1568.
Rattner, BA., Sileo, L., and Scanes, C.G. 1982a. Oviposition and the plasma concentrations of
LH, progesterone and corticosterone in bobwhite quail (Colinus virginianus) fed parathion.
Journal of Reproduction and Fertility 66: 147-155.
Rattner, BA., Sileo, L., and Scanes, C.G. 1982b. Hormonal responses and tolerance to cold of
female quail following parathion ingestion. Pesticide Biochemistry and Physiology 18:
132-138.
Rattner, B., Eroschenko, V., Fox, G., Fry, D., and Gorsline, J. 1984. Avian endocrine responses
to environmental pollutants. The Journal of Experimental Zoology 232: 683-689.
178
-------
Rattner, B.A., Hoffman, DJ., and Mam, C.M. 1989. Use of mixed-function oxygenases to
monitor contaminant exposure in wildlife. Environmental Toxicology and Chemistry 8:
1093-1102.
Reardon, C. and Lucas, D. 1987. Heavy-metal mitogenesis: Zn++ and Hg++ induce cellular
cytotoxicity and interferon production in murine T lymphocytes. Immunobiology 175(5):
455-469.
Redfield, A.C. 1958. The biological control of chemical factors in the environment. American
Scientist 46: 205-221.
Reeves, R. and Mitchell, E. 1984. Catch history and initial population of white whales
(Delphinapterus leucas) in the river and Gulf of St. Lawrence, Eastern Canada.
Naturaliste Canada (Review Ecology Systematics) 111: 63-121.
Rehana, T. and Rao, P.R. 1992. Effect of DDT on the immune system in Swiss Albino mice
during adult and perinatal exposure: humoral responses. Bulletin of Environmental
Contamination and Toxicology 48: 525-540.
Reijnders, P.J.H. 1986. Reproductive failure in common seals feeding on fish from polluted
waters. Nature 324: 456-457.
Reijnders, P. 1988. Environmental impact of PCBs in the marine environment. Pp. 86-98 in:
Newman, P.J. and Agg, A.R. (eds.). Environmental Protection of the North Sea. Oxford,
England: Heineman Professional Publishing.
Reijnders, P.J.H. and Brasseur, S.MJ.M. 1992. Xenobiotic induced hormonal and associated
developmental disordes in marine organisms and related effects in humans; an overview.
In: Colborn, T. and Clement, C. (eds.). Chemically-induced Alterations in Sexual and
Functional Development: The Human-Wildlife Connection. Princeton, NJ: Princeton
Scientific Publishing, Inc. In press.
Reyes, J., Reisz-Porszasz, S., and Hankinson, O. 1992. Identification of the Ah receptor nuclear
translocator protein (amt) as a component of the DNA binding form of the Ah receptor.
Science 256: 1193-1195.
Rice, C.P. and Evans, M.S. 1984. Toxaphene in the Great Lakes. Pp. 163-194 in: Nriagu, J.O.
and Simmons, M.S. (eds.). Toxic Contaminants in the Great Lakes. New York, NY: John
Wiley & Sons.
Rice, D.C. 1990. Delayed neurotoxicity in monkeys exposed developmentally to methylmercury.
Neurotoxicology 10: 645-50.
179
-------
Richie, PJ. and Peterle, TJ. 1979. Effect of DDE on circulation luteinizing hormone levels in
ring doves during courtship and nesting. Bulletin of Environmental Contamination and
Toxicology 23: 220-226.
Rickenbacher, U., McKinncy, J., Oatley, S., and Blake, C. 1986. Structurally specific binding of
halogenated biphenyls to thyroxine transport protein. Journal of Medical Chemistry 29:
641-648.
Riesbrough, R.W., Huggett, R., Grinnin, J., and Goldberg, E. 1968. Pesticides: Transatlantic
movements in the northeast trades. Science 159: 1233-1236.
Riznyk, R.Z., Hardy, J.T., Person, W., and Jabs, L. 1987. Short-term effects of polynuclear
aromatic hydrocarbons on sea-surface microlayerphytoneuston. Bulletin of Environmental
Contamination and Toxicology 38: 1037-1043.
Robblee, M.B., Barber, T.R., Carlson, P.R., Durako, M.J., Fourqurean, J.W., Muehlstein, L.K.,
Porter, D., Yarbro, L.A., Zieman, R.T., and Zieman, J.C. 1991. Mass mortality of the
tropical seagrass Thalassia testudinum in Florida Bay (US). Marine Ecology Progress
Series 71: 297-299.
Robineau, B., Gagne, J.A., Fortier, L., and Cembella, A.D. 1991. Potential impact of a toxic
dinoflagellate (Alexandrium excavation) bloom on survival of fish and crustacean larvae.
Marine Biology 108: 293-301.
Rodamilans, M., Osaba, M., To-Figueras, J., Fillat, F., Marques, J., Perez, P., and Corbella, J.
1988. Lead toxicity on endocrine testicular function in an occupationally exposed
population. Human Toxicology 7(2): 125-128.
Rodier, P.M., Ashmer, M., and Sager, P.R. 1984. Mitotic arrest in the developing CNS after
prenatal exposure to methylmercury. Neurobehavioral Toxicology and Teratology 6:379-
385.
Rogan, W., Gladen, B., McKinney, J., Carreras, N., Hardy, P., Thullen, J., Tingelstad, J., and
Tully, M. 1986a. Polychlorinated Wphenyls (PCBs) and dichlorodiphenyl
dichloroethene(DDE) in human milk: effects of maternal factors and previous lactation.
American Journal of Public Health 76: 172-177.
Rogan, W., Gladen, B., McKinney, J., Carreras, N., Hardy, P., Thullen, J., Tingelstad, J., and
Tully, M. 1986b. Neonatal effects of transplacental exposure to PCBs and DDE. The
Journal of Pediatrics 109: 335-341.
Rogan, W., Gladen, B., McKinney, J., Carreras, N., Hardy, P., Thullen, J., Tingelstad, J., and
Tully, M. 1987. Polychlorinated biphenyls (PCBs) and dichlorodiphenyl dichloroethene
180
-------
(DDE) in human milk: effects on growth, morbidity, and duration of lactation. American
Journal of Public Health 77: 1294-1297.
Rogan, W., Gladen, B., Hung, K, Koong, S., Shih, L., Taylor, J., Wu, Y., Yang, D., Ragan, N.,
and Hsu, C. 1988. Congenital poisoning by polychlorinated biphenyls and their
contaminants in Taiwan. Science 241: 334-336.
Romkes, M., Piskorska-Pliszczynska, J., and Safe, S. 1987. Effects of 2,3,7,8-
tetrachlorodibenzo-p-dioxin on hepatic and uterine estrogen receptor levels in rats.
Toxicology and Applied Pharmacology 87: 306-314.
Romkes, M. and Safe, S. 1988. Comparative activities of 2,3,7,8-tetrachlorodibenzo-p-dioxin
and progesterone as antiestrogens in the female rat uterus. Toxicology and Applied
Pharmacology 92(3): 368-380.
Rosenberg, R. 1985. Eutrophication — the future marine coastal nuisance? Marine Pollution
Bulletin 16: 227-231.
Rosenberg, R., Elmgren, R., Fleischer, S., Jonsson, P., Persson, G., and Dahlin, H. 1990. Marine
eutrophication case studies in Sweden. Ambio 19: 102-108.
Roth, W., Voonnan, R., and Aust, S. 1988. Activity of thyroid hormone-inducible enzymes
following treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicology and Applied
Pharmacology 92: 65-74.
Rounsefell, GA. and Dragovich, A. 1966. Correlation between oceanographic factors and
abundance of the Florida redtide (Gymnodinium breve Davis), 1954-1961. Bulletin of
Marine Science 16: 402.
Rourke, A.W., Eroschenko, V.P., and Washbum, L.J. 1991. Protein secretions in mouse uterus
after methoxychlor or estradiol exposure. Reproductive Toxicology 5(5): 437-442.
Rowe, G.T., Clifford, C.H., Smith, K.L., and Hamilton, P.L. 1975. Benthic nutrient regeneration
and its coupling to primary productivity in coastal waters. Nature 225: 215-217.
Rozman, K., D. Pereira, and M. latropoulos. 1987. Effect of a sublethal dose of 2,3,7,8-
tetrachlorodibenzo-p-dioxin on interscapular brown adipose tissue of rats. Toxicologic
Pathology 15(4): 425-430.
Rozman, K., Pfeifer, B., Kerecsen, L., and Alper, R.H. 1991. Is a serotonergic mechanism
involved in 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-induced appetite suppression
in the Sprague-Dawley rat? Archives of Toxicology 65(2): 124-128.
181
-------
Ruch, R., Klaunig, J., and Pereira, M. 1987. Inhibition of intercellular communication between
mouse hepatocytes by tumor promoters. Toxicology and Applied Pharmacology 87:111-
120.
Rudstam, L.G., Hansson, S., Johansson, S., and Larsson, U. 1992. Dynamics of planktivory in
a coastal area of the northern Baltic Sea. Marine Ecology Progress Series 80: 159-173.
Russell, D., Buckley, A., Shah, G., Sipes, L, Blask, D., and Benson, B. 1988. Hypothalamic site
of action of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicology and Applied
Pharmacology 94: 496-502.
Rydberg, L.L., Edler, S., Floderus, S., and Graneli, W. 1990. Interaction between supply of
nutrients, primary production, sedimentation and oxygen consumption in SE Kattegat.
Ambio 19: 134-141.
Ryther, J.H. 1954. The ecology of phytoplankton blooms in Moriches Bay and Great South Bay,
Long Island, New York. Biological Bulletin 106: 198-209.
Ryther, J.H. and Dunstan, W.M. 1971. Nitrogen, phosphorus and eutrophication in the coastal
marine environment. Science 171: 1008-1012.
Ryther, J.H. 1989. Historical perspective of phytoplankton blooms on Long Island and the green
tides of the 1950's. In: Cosper, E.M., Carpenter, EJ., and Bricelj, V.M. (eds.). Novel
Phytoplankton Blooms: Causes and Impacts of Recurrent Brown Tides and Other Unusual
Blooms. Lecture Notes on Coastal and Estuarine Studies. Berlin: Springer-Verlag.
Safe, S.H. 1984. Polychlorinated biphenyls (PCBs) and polybrominated biphenyls (PBBs):
biochemistry, toxicology and mechanism of action. CRC Critical Reviews of Toxicology
13(4): 319-395.
Safe, S.H., Bandiera, S., Sawyer, T., Okey, A., and Fujita, T. 1985. Effects of structure on
binding to the 2,3,7,8-TCDD receptor protein and AHH induction-halogeriated biphenyls.
Environmental Health Perspectives 61: 21-33.
Safe, S.H. 1986. Comparative toxicology and mechanism of action of polychlorinated dibenzo-
p-dioxins and dibenzofurans. Annual Review of Pharmacology and Toxicology 26: 371-
399.
Safe, S. 1987. Determination of 2,3,7,8-TCDD Toxic Equivalent Factors (TEFs): support for the
use of the in vitro AHH induction assay. Chemosphere 16: 791-802.
Safe, S. 1989. Polychlorinated biphenyls (PCBs): mutagenicity and carcinogenicity. Mutation
Research 220(1): 31-47.
182
-------
Safe, S., Astroff, B., Harris, M., Zacharewski, T., Dickerson, R., Romkes, M., and Biegel, L.
1991. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and related compounds as
antiestrogens: characterization and mechanism of action. Pharmacology and Toxicology
69(6): 400-409.
Sager, P.R., Doherty, RA., and Rodier, P.M. 1982. Morphometric analysis of the effect of
methylmercury on developing mouse cerebellar cortex. Toxicologist 2: 16.
Sager, D.B. 1983. Effect of postnatal exposure to polychlorinated biphenyls on adult male
reproductive function. Environmental Research 31: 76-94.
Sager, P.R., Doherty, RA., and Olmstead, J.B. 1983. Interaction of methylmercury with
microtubules in cultured cells and in vitro. Experimental Cell Research 146: 127-137.
Sager, P.R., Aschner, M., and Rodier, P.M. 1984. Persistent differential alterations in developing
cerebellar cortex of male and female mice after methylmercury exposure. Developmental
Brain Research 12: 1-11.
Sager, D., Shih-Schroeder, W., and Girard, D. 1987. Effect of early postnatal exposure to
polychlorinated biphenyls (PCBs) on fertility in male rats. Bulletin of Environmental
Contamination and Toxicology 38: 946-953.
Sarafian, T. and Verity, MA. 1985. Inhibition of RNA and protein synthesis in isolated
cerebellar cells by in vitro and in vivo methylmercury. Neurochemical Pathology 3:27-39.
Sarafian, T. and Verity, MA. 1986. Mechanism of apparent transcription inhibition by
methyimercury in cerebellar neurons. Journal of Neurochemistry 47: 625-631.
Sarokin, D. and Schulkin. J. 1992. The role of pollution in large scale population disturbances,
Part 1: Aquatic. Environmental Science and Technology 26(8): 1476-1484.
Sawyer, T.W., Vatcher, A.D., and Safe, S. 1984. Comparative aryl hydrocarbon hydroxylase
induction activities of commercial PCBs in Wistar rates and rate hepatoma H-4-IIE cells
in culture. Chemosphere 13: 695-701.
Saxena, M.C., Siddiqui, M.K.J., Agarwal, V., and Kutty, D. 1983. A comparison of
organochlorine insecticide contents in specimens of maternal blood, placenta, and
umbilical cord-blood from still-born and live-born cases. Journal of Toxicology and
Environmental Health 11: 71-79.
Saxena, M.P:, Gopal, K., Jones, W., and Ray, P.K. 1992. Immune responses to Aeromonas
hydrophila in cat fish (Heteropneustis fossilis) exposed to cadmium and
hexachlorocyclohexane. Bulletin of Environmental Contamination and Toxicology 48:
194-201.
183
-------
Schantz, S.L., Barsotti, DA., and Allen, J.R. 1979. Toxicological effects produced in non-human
primates chronically exposed to fifty parts per trillion 2,3,7,8-tetrachlorodibenzo-p-
dioxin (TCDD). Toxicology and Applied Pharmacology (Part 2) 48: A180.
Schantz, S.L. and Bowman, R.E. 1989. Learning in monkeys exposed perinatally to 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD). Neurotoxicology and Teratology 11: 13-19.
Schecter, A., Mes, J., and Davies, D. 1989. Polychlorinated biphenyl (PCB), DDT, DDE and
hexachlorobenzene (HCB) and PCDD/F isomer levels in various organs in autopsy tissue
from North American patients. Chemosphere 18(1-6): 812-818.
Schecter, A., Papke, O., and Ball M. 1990. Evidence for transplacental transfer of dioxins from
mother to fetus: chlorinated dioxin and dibenzofuran levels in the livers of stillborn
infants. Chemosphere 21(8): 1017-1022.
Schecter, A., McGee, H., Stanley, J., and Boggess, K. 1992. Chlorinated dioxin, dibenzofuran,
coplanar, mono-orthoi and di-ortho substituted PCB congener levels in blood and semen
of Michigan Vietnam veterans compared with levels in Vietnamese exposed to agent
orange. Submitted to Chemosphere November 1992. In press.
Schelske, C.L. and Hodell, D.A. 1991. Recent changes hi productivity and climate of Lake
Ontario detected by isotopic analysis of sediments. Limnology & Oceanography 36: 961-
975.
Schiller, C.M., Walden, R., and Shoaf, C.R. 1982. Studies on the mechanism of 2,3,7,8-
tetrachlorodibenzo-p-dioxin toxicity: Nutrient assimilation. Federal Procedure 41: 1426.
(Abst).
Schindler, D.W. 1977. Evolution of phosphorus limitation in lakes. Science .195: 260-262.
Schindler, D.W., Hesslein, R., and Kipphut, G. 1977. Interactions between sediments and
overlying waters in an experimentally eutrophied Pre-Cambrian shield lake. In:
Goltterman, H.L. (ed.). Interactions Between Sediments and Fresh Water. Junk, The
Hague.
Schindler, D.W., Fee, E.S., and Roszcynski, T. 1978. Phosphorus input and its consequences for
phytoplankton standing crop and production in the Experimental Lakes Area and in
similar lakes. Journal of the Fisheries Research Board of Canada 35: 190-196.
Schindler, D.W. 1978. Factors regulating phytoplankton production and standing crop in the
world's freshwaters.- Limnology & Oceanography 23: 478-486.
184
-------
Schindler, D.W. 1981. Studies of eutrophication in lakes and their relevance to the estuarine
environment. In: Neilson, BJ. and Cronin, L.E. (eds.). Estuaries and Nutrients. Humana,
NY.
Schindler, D.W., Mills, K.H., Mailey, D.F., Findlay, D.L., Shearer, J.A., Davies, I.J., Turner,
M.A., Linsey, GA., and Cruikshank, D.R. 1985. Long-term ecosystem stress; the effects
of years of experimental acification on a small lake. Science 228: 1395-1401.
Schindler, D.W., 1987. Determining ecosystem responses to anthropogenic stress. Canada Journal
of Fisheries and Aquatic Science 44 (supp. 1): 6-25.
Schmidt, K.F. 1992. Dioxin's other face: portrait of an "environmental hormone". Science News
141: 24-27.
Schmitt, C.J., Zajicek, J.L., and Ribick, M.A. 1985. National Pesticide Monitoring Program:
residues of organochlorine chemicals in fresh water fish, 1980-81. Archives of
Environmental Contamination and Toxicology 14: 225-260.
Schmitt, C.J., Ludke, J.L., and Walsh, D. 1981. Organochlorine residues in freshwater fish,
1976-1979; National Pesticide Monitoring Program. Pesticides Monitoring Journal 14:
136-206.
Schrenk, D., Karger, A., Lipp, H.P., and Bock, K.W. 1992. 2,3,7,8-Tetrachlorodibenzo-p-dioxin
and ethinylestradiol as co-mitogens in cultured rat hepatocytes. Carcinogenesis 13(3):
453-456.
Schwartz, J., Jacobson, S., Fern, G., Jacobson, J., and Price, H. 1983. Lake Michigan fish
consumption as a source of polychlorinated biphenyls in human cord serum, maternal
serum, and milk. American Journal of Public Health 73(3): 293-296.
Scott, B.C. 1981. Modeling of atmospheric wet deposition. Pp. 3-21 in: Eisenreich, S.J. (ed.).
Atmospheric Inputs of Pollutants to Natural Waters. Ann Arbor, MI: Science Publishers.
Seba, D.B. and Prospero, J.M. 1971. Pesticides in the lower atmosphere of the northern
equatorial Atlantic Ocean. Atmospheric Environment 5: 1043-1050.
Seba, D.B. and Prospero, J.M. 1972. Some additional measurements of pesticides in the lower
atmosphere of the northern equatorial Atlantic Ocean. Atmospheric Environment 6: 363-
364.
Seegal, R.F., Brosch, K.O., and Bush, B. 1985. Oral dosing of rats with polychlorinated
biphenyls increases urinary homovanillic acid production. Journal of Toxicology and
Environmental Health 15: 575-586.
185
-------
Seegal, R.F., Brosch, K.O., and Okoniewski, R. 1988. The degree of PCB chlorination determines
whether the rise in urinary homovanillic acid production in rats is peripheral or central
in origin. Toxicology and Applied Pharmacology 96(3): 560-564.
Seegal, R., Bush,B., and Shain, W. 1990. Lightly chlorinated ortho-substituted PCB congeners
decrease dopamine in nonhuman primate brain and in tissue culture. Toxicology and
Applied Pharmacology 106(1): 136-144.
Seegal, R.F., Bush, B., and Brosch, K.0.1991a. Subchronic exposure of the adult rat to Aroclor
1254 yields regionally-specific changes in central dopaminergic function.
Neurotoxicology 12(1): 55-65.
Seegal, R.F., Bush, B., and Brosch, K.O. 1991b. Comparison of effects of Aroclors 1016 and
1260 on non-human primate catecholamine function. Toxicology 66(2): 145-163.
Seegal, R.F. 1992a. Study in progress. Wadsworth Center, New York State Department of Health
of Environmental Health and Toxicology, School of Public Health, University of Albany.
Albany, NY.
Seegal, R.F. 1992b. Perinatal exposure to arochlor 1016 elevates brain dopamine concentrations
in the rat. Journal of Toxicology. In press.
Seitzinger, S.P. 1988. Denitrification in freshwater and marine ecosystems: ecological and
geochemical significance. Limnology & Oceanography 33: 702-724.
Seitzinger, S.P., Gardner, W.S., and Spratt, A.K. 1991. The effect of salinity on ammonium
sorption in aquatic sediments: implications for benthic nutrient cycling. Estuaries 14:167-
174.
Selvan, R.S., T.N. Dean, H.P. Misra, P.S. Nagarkatti, and M. Nagarkatti. 1989. Andicarb
suppresses macrophage but not natural killer (NK) cell-mediated cytotoxicity of tumor
cells. Bulletin of Enviromental Contamination and Toxicology 43: 676-682.
Sergeant, D. 1986. Present status of white whales (Delphinapterus leucas) in the St. Lawrence
Estuary. Naturaliste Canada (Reviews Ecology Systematics) 113: 61-81.
Setzler-Hamilton, E.M., Whipple, J.A., and MacFariane, R.B. 1988. Striped bass populations in
Chesapeake and San Francisco Bays: two environmentally impacted estuaries. Marine
Pollution Bulletin 19(9): 466-477.
Shain, W., Seegal, R., Priester, K., and Bush, B. 1990. Structure/activity relationship for PCB
neurotoxicity. Paper No. 404 at the SETAC Annual Meeting, Global Environmental
Issues: Challenges for the 90's. Arlington, VA.
186
-------
Shain, W., Bush, B., and Seegal, R. 1991. Neurotoxicity of polychlorinated biphenyls: structure-
activity relationship of individual congeners. Toxicology and Applied Pharmacology
111(1): 33-42.
Sharpe, R.M. 1992. Declining sperm counts in men — is there an estrogen cause? Journal of
Endocrinology. In press.
Shimai, S. and Satoh, H. 1985. Behavioral teratology of methylmercury. Journal of Toxicological
Sciences 10: 199-216.
Shoaf, C.R. and Schiller, C.M. 1981. Studies on the mechanism of 2,3,7,8-tetrachlorodibenzo-p-
dioxin (TCDD) toxicity-lipid assimilation, n. Pharmacologist 23: 176. (Abstr).
Short, F.T., Davis, M.W., Gibson, R.A., and Zimmerman, CF. 1985. Evidence for phosphorus
limitation in carbonate sediments of the seagrass Syringodium filiforme. Estuarian and
Coastal Shelf Science 20: 419-430.
Short, F.T., Dennison, W.C., and Cappone, D.G. 1990. Phosphorus-limited growth of the tropical
seagrass Syringodium filiforme in carbonate sediments. Marine Ecology Progress Series
62: 169-174.
Shugart, G. 1980. Frequency and distribution of polygony in Great Lakes herring gulls in 1978.
Condor 82: 426-429.
Shugart, G., Fitch, M.A., and Fox, G A. 1988. Female pairing: a reproductive strategy for herring
gulls. The Condor 90: 933-935.
Sieburth, J.P., Johnson, W., and Hargraves, P.E. 1988. Ultrastructure and ecology of Aureococcus
anophagefferens gen. et sp. nov. (Chrysophyceae); the dominant picoplankter during a
bloom in Narragansett Bay, Rhode Island, Summer 1985. Journal of Phycology 24: 416-
425.
Silbergeld, E. and Mattison, D. 1987. Experimental and clinical studies on the reproductive
toxicology of 2,3,7,8-tetrachlorodibenzo-p-dioxin. American Journal of Industrial
Medicine 11(2): 131-144.
Sileo, L., Karstad, L., Frank, R., Holdrinet, M., Addison, E., and H. Braun. 1977. Organochlorine
poisoning of ring-billed gulls in Southern Ontario. Journal of Wildlife Diseases 13: 313-
322.
Simic, B., Kniewald, Z., Davies, J.E., and Kniewald, J. 1991. Reversibility of the inhibitory
effect of atrazine and lindane on cytosol 15 alpha-dihydrotestosterone. Bulletin of
Environmental Contamination and Toxicology 46: 92-99.
187
-------
Simmonds, M. 1991. Cetacean mass mortalities and their potential relationship with pollution.
The Symposium on Whales-Biology-Threats-Conservation. Brussels. June 5-7.
Simpson, J.G. and Gardner, M.B. 1972. Comparative anatomy of selected marine mammals. Pp.
298-418 in: Ridgway, S.H. (ed.). Mammals of the Sea: Biology and Medicine.
Springfield, IL: CC Thomas.
Simpson, E.R. and Waterman, M.R. 1989. Steroid hormone biosynthesis in the adrenal cortex and
its regulation by adrenocorticotropin. Pp. 1543-1556 in: DeGroot, L.R. (ed.).
Endocrinology, Volume 3, 2nd Edition. Philadelphia, PA: W.B. Saunders Co.
Singhal, R.L. Valadares, J.R.E., and Schwark, W.S. 1970. Metabolic control mechanism in
mammalian systems. DC. Estrogen-like stimulation of uterine enzymes by o,p'-l,l,l,-
trichloro-2-2-bis(p-chlorophcnyl)ethane. Biochemical Pharmacology 19: 21245-2155.
Slinn, S.A. and Slinn, W.G.N. 1980. Prediction for particle deposition on natural waters.
Atmospheric Environment 14: 1013-1016.
Slinn, W.G.N., Hasse, L., Hicks, B., Hogan, A., Lai, D., Liss, P., Munnich, K., Sehmel, G., and
Vittori, O. 1978. Some aspects of the transfer of atmospheric trace constituents past the
AIR-SEA interface. Atmospheric Environment 12: 2055-2087.
Sloof, W. and Matthijsen, A. 1988. Integrated Criteria Document Hexachlorocyclohexanes.
Report No. 758473011. National Institute of Public Health and Environmental Protection,
Bilthoven, The Netherlands. October.
Slotkin, T.A., Pachman, S., Kazlock, RJ., and Bartolome, J. 1985. Effects of neonatal
methylmercury exposure on development of nucleic acids and proteins in rat brain:
regional specificity. Brain Research Bulletin 14: 397-400.
Smayda, T.J. 1974. Bioassay of the growth potential of the surface water of lower Narragansett
Bay over an annual cycle using the diatom Thalassiosira pseudonana (oceanic clone, 13-
1). Limnology & Oceanography 19: 889-901.
Smayda, TJ. 1992. A phantom of the ocean. Nature 358: 374-375.
Smialowicz, RJ., Andrews, J.E., Riddle, M.M., Rogers, R.R., Luebke, R.W., and Copeland, C.B.
1989. Evaluation of the immunotoxicity of low level PCS exposure in the rat. Toxicology
56(2): 197-211.
Smith, S.V. 1981. Responses of Kaneohe Bay, Hawaii, to relaxation of sewage stress. In:
Neilson, J. and Cronin, L.E. (eds.). Estuaries and Nutrients. Humana, NY.
188
-------
Smith, S.V. 1984. Phosphorus vs. nitrogen limitation in the marine environment. Limnology &
Oceanography 29: 1149-1160.
Smith, S.V. and Atkinson, M.J. 1984. Phosphorus limitation of net production in a confined
aquatic ecosystem. Nature 207: 626-627.
Smith, V.H. 1979. Nutrient dependence of primary productivity in lakes. Limnology &
Oceanography 24: 1051-1064.
Smith, V.H. 1990. Nitrogen, phosphorus, an nitrogen fixation in lacustrine and estuarine
ecosystems. Limnology & Oceanography 35: 1852-1859.
Sodergren, A. and Gelin, C 1983. Effect of PCBs on the rate of carbon-14 uptake in
phytoplankton isolates from oligotrophic and eutrophic lakes. Bulletin of Environmental
Contamination and Toxicology 30: 191-198.
Sonawane, B., Smialowicz, R., and Luebke, R. 1988. Immunotoxicity of 2,3,7,8-TCDD: review,
issues, and uncertainties. Appendix E. In: U.S. Environmental Protection Agency. A
Cancer Risk-Specific Dose Estimate for 2,3,7,8-TCDD (Review Draft) (Appendices A
through F). Office of Health and Environmental Assessment. EPA/600/6-88/007Ab.
Sonzogni, W.C. and Swain, W.R. 1980. Perspectives on U.S. Great Lakes chemical toxic
substances research. Journal of Great Lakes Research 6: 265-274.
Spear, P.A. and Moon, T.W. 1985 Low dietary iodine and thyroid anomalies in ring doves,
Streptopelia risoris, exposed to 3,4,3'4'-tetrachlorobiphenyl. Archives of Environmental
Contamination and Toxicology 14: 547-553.
Spencer, D., House, I., Tripp, J., and Stimmler, L. 1988. Mercury concentration in cord blood.
Archives of Disease in Childhood 63: 202-203. - •
Spencer, W.F. 1974. Movement of DDT and its derivatives into the atmosphere. Research
Review 59: 91-117.
Spies, R.B. Rice, D.W., and Ireland, R.R. 1984. Preliminary studies of growth, reproduction and
activity of hepatic mixes-function oxidase \nPlatichthysstellatus. Marine Environmental
Research 14: 426-428.
Spink, D.C., Lincoln, D.W. 0, Dickerman, H.W., and Gierthy, J.F. 1990. 2,3,7,8-
tetrachlorodibenzo-p-dioxin causes an extensive alteration of 17 beta-estradiol
metabolism in MCF-7 breast rumor cells. Proceedings of the National Academy of
Science, U.SA. 87(17): 6917-6921.
189
-------
Spink, D.C., Eugster, H.P., Lincoln, D.W. n, Schuetz, J.D., Schuetz, E.G., Johnson, J.A.,
Kaminsky, L.S., and Gierthy, J.F. 1992. 17 beta-estradiol hydroxylation catalyzed by
human cytochromc P450 1A1: a comparison of the activities induced by 2,3,7,8-
tetrachlorodibenzo-p-dioxin in MCF-7 cells with those from heterologous expression of
the DNA. Archives of Biochemistry and Biophysics 203(2): 342-348.
Spitsbergen, J.M., Kleeman, J.M., and Peterson, R.E. 1988. Morphologic lesions and acute
toxicity in rainbow trout (Salmo gairdneri) with 2,3,7,8-tetrachlorodibenzo-p-dioxin.
Journal of Toxicology and Environmental Health 23: 333-358.
Spyker, J.M., Sparber, S.B., and Goldberb, A.M. 1972. Subtle consequences of methylmercury
exposure: behavioural deviations in offspring of treated mothers. Science 177: 621-623.
Stahl, B.U. and Rozman, K. 1990. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)-induced
appetite suppression in the Sprague-Dawley rat is not a direct effect on feed intake
regulation in the brain. Toxicology and Applied Pharmacology 106(1): 158-162.
r
Stahl, B.U., Alper, R.H., and Rozman, K. 1991. Depletion of brain serotonin does not alter
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-induced starvation syndrome in the rat.
Toxicology letters 59: 65-72.
Stancel, G.M., Ireland, J.S., Mukku, V.R., and Robison, A.K. 1980. The estrogenic activity of
DDT: in vivo and in vitro induction of a specific estrogen inducible uterine protein by
o.p'DDT. Life Science 27: 1111-1117.
Steele, J.H. 1974. The Structure of Marine Ecosystems. Cambridge, MA: Harvard
University Press.
Sternowsky, H. and Wessolowski, R. 1985. Lead and cadmium in breast milk — higher levels
in urban vs. rural mothers during the "first 3 months of lactation. Archives of Toxicology
57: 41-45.
Stewart, F. and Smith, A. 1986. Metabolism of hexachlorobenzene by rat-liver microsomes. Pp.
325-327 in: Morris and Cabral (eds.). Hexachlorobenzene: Proceedings of an
International Symposium. IARC. Lyon, France.
Stohs, S.J., Abbot, B.D., Lin, F.H., and Bimbaum, L.S. 1990. Induction of ethoxyresomfin-O-
deethylase and inhibition of glucocorticoid receptor binding in liver of haired and hairless
HRS/J mice by topically applied 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicology 65:
123-136.
Strachan, S.M.J. and Huneault, H. 1979. Polychlorinated biphenyls and organochlorine pesticides
in Great Lakes precipitation. Journal of Great Lakes Research 5: 61-68.
190
-------
Strachan, W.MJ. and Eisenreich, SJ. 1988. Mass balancing of toxic chemicals in the Great
Lakes: the role of atmospheric deposition. Workshop Proceedings, Scarborough, Ontario.
International Joint Commission. November, 1986.
Streissguth, A.P., Landesman-Dwyer, S., Martin, J.C., and Smith, D.W. 1980. Teratogenic effects
of alcohol in humans and laboratory animals. Science 209: 353-361.
Streissguth, A.P., Barr, H.M., and Martin, D.C. 1983. Maternal alcohol use and neonatal
habituation assessed with the Brazelton scale. Child Development 54: 1109-1118.
Stressguth, A.P., Martin, D.C., Barr, H.M., Sandman, B.M., Kirchner, G.L., and Darby, B.L.
1984. Intrauterine alcohol and nicotine exposure: attention and reaction time in 4-year-
old children. Developmental Psychology 20: 533-541.
Subranianian, A.N., Tanabe, S., Tatsukawa, R., Saito, S., and Miyazaki, N. 1987. Reduction in
the testosterone levels by PCBs and DDE in Dall's porpoises of Northerwestern North
Pacific. Marine Pollution Bulletin 18(12): 643-646.
Sukumar, A. and Karpagaganapathy, P.R. 1992. Pesticide-induced atresia in ovary of fresh water
fish (Colisa alia). Bulletin of Contamination and Toxicology 48: 457-462.
Suresh, A., Sivaramakrishna, B., Victoriamma, P.C., and Radhakrishnaiah, K. 1992. Comparative
study on the inhibition of acetylcholinesterase activity in the freshwater fish Cyprinus
carpio by mercury and zinc. Biochemistry International 26: 367-375.
Swackhamer, D.L. and Kites, RA. 1988. Occurrence and bioaccumulation of organochlorine
compounds in fishes from Siskiwit Lake, Isle Royale. Environmental Science and
Technology 22: 543-548.
Swackhamer, D.L., McVeety, B.V., and Kites, RA. 1988. Deposition and evaporation of PCB
congeners to and from Siskiwit Lake, Isle Royale. Environmental Science and Technology
22: 664-672.
Swackhamer, D.L. and Armstrong, D.E. 1988. Horizontal and vertical distribution of PCBs in
southern Lake Michigan sediments and the effect of Waukegan as a point source. Journal
of Great Lakes Research 14: 277-290.
Swackhamer, D.L., Pearson, R., and Holmes, M. 1992. Unpublished data, University of
Minnesota.
Swain, W.R. 1978. Chlorinated organic residues in fish, water and precipitation from the vicinity
of Isle Royale, Lake Superior. Journal of Great Lakes Research 4: 398-407.
191
-------
Swain, W.R., Mullin, M.D., and Filkins, J.C. 1986. Long range transport of toxic organic
contaminants to the North American Great Lakes. Pp. 107-121 in: Problems of aquatic
Toxicology, Biotesting, and Water Quality Management: Proceedings of USA-USSR
Symposium, Barak, Jaroslavl Oblast, July 30-August 1, 1984. U.S. Environmental
Protection Agency; EPA/600/9-86/024.
Swain, W.R. 1988a. Human health consequences of consumption of fish contaminated with
organochlorine compounds. Aquatic Toxicology 11: 357-377.
Swain, W.R. 1988b. Evidence of long-range atmospheric transport of toxic xenobiotic substances
on the Great Lakes region. Testimony before the Subcommittee on Investigations and
Oversight of the Committee on Public Works and Transportation, U.S. House of
Representatives. Hearing on Long Range Transport of Toxic Chemicals to the Great
Lakes. April 14.
Szmcynski, G. and Waliszewski, S. 1981. Chlorinated pesticide residues in testicular tissue
samples, pesticides in human testicles. Andrologia 15(6): 696-698.
Takeuchi, T. 1972a. Approaches to the detection of subclinical mercury intoxications: experience
in Minimata, Japan. In: Hartung, R. and Dinman, B.D. (eds.). Environmental mercury
contamination. Ann Arbor, MI: Science Press.
Takeuchi, T. 1972b. Biological reactions and pathological changes in human beings and animals
caused by organic mercury contamination. Pp. 82-96 in: Hartung, R. and Dinman, B.D.
(eds.). Environmental Mercury Contamination. Ann Arbor, MI: Science Press.
Tanabe, S., Kannan, N., Subramanian, A., Watanabe, S., and Tatsukawa, R. 1987. Highly toxic
coplanar PCBs: occurrence, source, persistency and toxic implications to wildlife and
humans. Environmental Pollution 47: 147-163.
Thakker, D.R., Yagi, H., Levin, W., Wood, A.W., Conney, A.H., and Jerina, D.M. 1985.
Polycyclic aromatic hydrocarbons: metabolic activation to ultimate carcinogens. Pp. 178-
242 in: Anders, M.W. (ed.). Bioactivation of Foreign Compounds. New York, NY:
Academic Press.
Thiyagarajah, A., Zwemer, D.E., and Hargis, Jr., WJ. 1989. Renal lesions in estuarine fishes
collected from the Elizabeth River, Virginia. Journal of Environmental Pathology,
Toxicology, and Oncology 9: 261-268.
Thomann, R.V. and Connolly, J.P. 1984. Model of PCB in the Lake Michigan lake trout food
chain. Environmental Science Technology 18: 65-72.
192
-------
Thomas, DJ. and Syversen, T.L.M. 1987. The alteration of protein synthesis by methyl mercury.
Pp. 131-171 in: Eccles, CU. and Annau, Z. (eds.). The Toxicity of Methyl Mercury.
Baltimore, MA: Johns Hopkins University Press.
Thomas, P. 1988. Reproductive endocrine function in female Atlantic craoker exposed to
pollutants. Marine Environmental Research 24: 179-183.
Tillitt, D.E., Ankley, G.T., Giesy, J.P., Kevern, N.R. 1988a. The use of H4IIE rat hepatoma cell
assay for the calculation of 2,3,7,8-Tetrachlorodibenzo-p-dioxin equivalents in
environmental samples. Report to U.S. Fish and Wildlife Service. Cooperative Agreement
14-16-003-87-943.
Tillitt, D.E., Ankley, G.T., Giesy, J.P., Kevem, N.R. 1988b. H4IIE rat hepatoma cell extract
bioassay-derived 2,3,7,8-Tetrachlorodibenzo-p-dioxin-quivalents (TCDD-EQ) from
Michigan waterbird colony eggs 1986 and 1987. Pesticide Research Center Report,
Michigan State University, East Lansing, MI.
Tillitt. D.E., Ankley, G.T., Giesy, J.P., Ludwig, J.P., Kurita-Matsuba, H., Wesehloh, D.V., Ross,
P.S., Bishop, CA., Sileo, L., Stromborg, K.L., Larson, J., and Kubiak, TJ. 1992.
Polychlorinated biphenyl residues and egg mortality in double crested cormorants from
the Great Lakes. Environmetnal Toxicology and Chemistry 11: 1281-1288.
Tilson, H.A., Jacobson, J.L., and Rogan, WJ. 1990. Polychlorinated biphenyls and the
developing nervous system: cross-species comparisons. Neurotoxicology and Teratology
12: 239-248.
Tomar, R.S. and Kerkvliet, N.I. 1991. Reduced T-helper cell function in mice exposed to
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicology Letters 57(1): 55-64.
Traber, P., Chianale, J., Florence, R., Kim, K., Wojcik, E., and Gumucio, J. 1988. Expression of
cytochrome P450b and P450e genes in small intestinal mucosa of rats following treatment
with phenobarbital, polyhalogenated biphenyls, and organochlorine pesticides. The Journal
of Biological Chemistry 263(19): 9449-9455.
Tracey, GA., Steele, R.L., Gatzke, J., Phelps, D.K., Nuzzi, R., Waters, M., and Anderson, D.M.
1989. Testing and application of biomonitoring methods for assessing environmental
effects of noxious algal blooms. In: Cosper, E.M., Carpenter, E.J., and Bricelj, V.M.
(eds.). Novel Phytoplankton Blooms: Causes and Impacts of Recurrent Brown Tides and
Other Unusual Blooms. Lecture Notes on Coastal and Estuarine Studies. Berlin: Springer-
Verlag.
Trapp, M., Baukloh, V., Bonnet, H.G. 1984. Pollutants in human follicular fluid. Fertility and
Sterility 42: 146-148.
193
-------
Trosko, J., and Chang, C. Non-genotoxic mechanisms in carcinogencsis: role of inhibited
intercellular communication. Branbury Report. In press.
Truelove, J.F., Tanner, J.R., Langlois, LA., Stapley, RA., Arnold., D.L., and Mes, J.C. 1990.
Effect of polychlorinated biphenyls on several endocrine reproductive parameters in the
female rhesus monkey. Archives of Environmental Contamination and Toxicology 19(6):
939-943.
Truong, D.D., Garcia De Yebenes, J., Pezzoli, G., Jackson-Lewis, V., and Fahn, S. 1988.
Glycine involvement in DDT-induced myoclonus. Movement Disorders 3(1): 77-87.
Truscott, B., Walsh, J.M., Burton, M.P., Payne, J.F., and Idler, D.R. 1983. Effect of acute
exposure to crude petroleum on some reproductive hormones in salmon and flounder.
Comparative Biochemistry and Physiology 75C: 121-130.
Tryphonas, H., Hayward, S., O'Grady, L., Loo, J.C., Arnold, D.L., Bryce, F., and Zawidzka, Z.Z.
1989. Immunotoxicity studies of PCB (Aroclor) 1254 in the adult rhesus (Macaco
mulatto) monkey — preliminary report. International Journal of Immunopharmacology
11(2): 199-206.
Tryphonas, H., Luster, M.I., White, K.L. Jr., Naylor, P.H., Erdos, M.R., Burleson, G.R.,
Gennolec, D., Hodgen, M., Hayward, S., and Arnold, D.L. 1991a. Effects of PCB
(Aroclor 1254) on non-specific immune parameters in rhesus (Macaco mulatto) monkeys.
International Journal of Immunopharmacology 13(6): 639-648.
Tryphonas, H., Luster, M.I., Schiffman, G., Dawson, L.L., Hodgen, M., Gennolec, D., Hayward,
S., Bryce, F., Loo, J.C.K., and Mandy, F. 1991b. Effect of chronic exposure of PCB
(Aroclor 1254) on specific and nonspecific immune parameters in the rhesus (Macaco
mulatto) monkey. Fundamentals of Applied Toxicology 16(4): 773-380.
Tuchmann-Duplessis, H. 1975. Drug effects on the fetus. Monographs on Drugs, Vol. II. Sydney,
Australia. ADIS Press.
Tuomisto, J., Pohjanvirta, R., MacDonald, E., and Tuomisto, L. 1990. Changes in rat brain
monoamines, monoamine metabolites and histamine after a single administration of
2,3,7,8-tetrachlorodibenzo-p-dioxin(TCDD). Pharmacology and Toxicology 67(3): 260-
265.
Tuppurainen. M., Wagar, G., Kurppa, K., Sakari, W., Wambugu, A., Froseth, B., Alho, J., and
Nykyri, E. 1988. Thyroid function as assessed by routine laboratory tests of workers with
long-term lead exposure. Scandinavian Journal of Work, Environment and Health 14(3):
175-180.
194
-------
Twilley, R.R., Kemp, W.M., Staver, K.W., Stevenson, J.C., and Boynton, W.R. 1985. Nutrient
enrichment of estuarine submerged vascular plant communities. 1. Algal growth and
effects on production of plants and associated communities. Marine Ecology Progress
Series 23: 179-191.
Umbach, J., Boadi, W., Brandes, J.M., Deraer, H., and Yannai, S. 1992. Effect of inorganic
mercury on in vitro placental nutrient transfer and oxygen consumption. Reproductive
Toxicology 6: 69-75.
Umbreit, T.H. and Gallo, M. 1988. Physiological implications of estrogen receptor modulation
by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicology Letters 42(1): 5-14.
Umbreit, T.H., Hesse, E.J., MacDonald, GJ., and Gallo, MA. 1988. Effects of TCDD-estradiol
interactions in three strains of mice. Toxicology Letters 40: 1-9.
Umbreit, T.H., Engles, D., Grossman, A., and Gallo, M.A. 1989a. Species comparison of steroid
UDP-glucuronyl transferase: correlation to TCDD sensitivity. Toxicology Letters 48(1):
29-34.
Umbreit, T.H., Scala, P.L., MacKenzie, S.A., and Gallo, M.A. 1989b. Alteration of the acute
toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) by estradiol and tomoxifen.
Toxicology 1989 59(2): 163-169.
United Nations Environmental Programme (UNEP). 1991. Review of Contaminants in Marine
Mammals. UNEP Marine Mammal Technical Report Number 2, ICES/IOC/UNEP,
Nairobi.
United States Environmental Protection Agency (USEPA). 1971. Pollution of the interstate waters
of Long Island Sound and its tributaries — Connecticut-New York. Washington, DC.
Government Printing Office.
United States Environmental Protection Agency (USEPA). 1985. Drinking Water Criteria
Document for 2,3,7,8-Tetrachlorodibenzo-p-Dioxin. EPA-440/5-84-007.
United States Environmental Protection Agency (USEPA). 1987. Hexachlorobenzene. Health
Advisory Draft. Office of Drinking Water. March 31, 1987.
United States Environmental Protection Agency (USEPA). 1987. Mercury. Health Advisory
Draft. Office of Drinking Water. March 31, 1987.
United States Environmental Protection Agency (USEPA). September 1991. Preliminary Draft:
Health Assessment for 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) And Related
Compounds. Environmental Criteria and Assessment Office. Cincinnati, Ohio.
195
-------
United States Environmental Protection Agency (USEPA). 1991. Long Island Sound Study.
Status Report and Interim Actions for Hypoxia Management. Environmental Protection
Agency. Draft Report.
United States Environmental Protection Agency (USEPA). 1991. Toxics in the Community: the
1989 Toxics Release Inventory National Report. U.S. EPA Office of Toxic Substances.
Economics and Technology Division. Washington, DC.
United States National Human Adipose Tissue Survey.
United States Public Health Service. 1988. ATSDR. Toxicological Profile for Lead (Draft). Oak
Ridge National Laboratory. February.
Uphouse, L. 1987. Decreased rodent sexual receptivity after lindane. Toxicology Letters 42(1):
5-14.
Uphouse, L. and Williams, J. 1989. Sexual behavior of intact female rats after treatment with
ojp'-DDT or p,p'-DDT. Reproductive Toxicology 3(1): 33-41.
Uphouse, L., Eckols, K., Croissant, D., and Stewart, G. 1990. Serotonergic changes following
proestrous treatment with p,p'-DDT. Neurotoxicology 11(3): 533-538.
Valiela, I. 1984. Marine Ecological Processes. New York, NY: Springer-Verlag.
Varanasi, U., Chan, S-L., McCain, B.B., Landahl, J.T., Schiewe, M.H., Clark, Jr., R.C., Brown,
D.W., Myers, M.S., Krahn, M.M., Gronlund, W.D., and MacLeod, Jr., W.W. 1989.
National Benthic Surveillance Project: Pacific Coast, Part n, Technical Presentation of the
Results for Cycles I to III (1984-1986). NOAA Technical Memo. NMFS F/NWC-170.
Vecchi, A., Mantovani, A., Sironi, M., Luini, W., Spreafico, F., and Garattini, S. 1980.
Immunosuppressive effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on humoral
antibody production and cell-mediated activities in mice. Archive of Toxicology 4:163-
165.
Veith, G.D., Kuehl, D.W., Puglisi, F.A., Glass, G.E., and Eaton, J.G. 1977. Residues of PCBs
and DDT in the western Lake Superior ecosystem. Archives of Environmental
Contamination and Toxicology 5: 487-499.
van Velsen, F., Danse, L., van Leeuwen, F., Dormans, J., and van Logten, M. 1986. The
subchronic oral toxicity of the b-isomer of hexachlorocyclohexane in rats. Fundamental
and Applied Toxicology 6: 697-712.
196
-------
Veltman, J. and Maines, M. 1986. Alterations of heme, cytochromc p-450, and steroid
metabolism by mercury in rat adrenal. Archives of Biochemistry and Biophysics 248(2):
467-478.
Verschaeve, L. and Leonard, A. 1984. Dominant lethal test in female mice treated with
methylmercury chloride. Mutation Research 136: 131-136.
Verschueren, K. 1983. Handbook of Environmental Data on Organic Chemicals. 2nd Edition.
New York, NY: Van Nostrand Reinhold Company.
Vince, S. and Valiela, I. 1973. The effects of ammonium and phosphate enrichment on
chlorophyll, a pigment ratio and species composition of phytoplankton of Vineyard Sound.
Marine Biology 19: 69-73.
Vitousek, P.M. and Howarth, R.W. 1991. Nitrogen limitation on land and in the sea: how can
it occur? Biogeochemistry 13: 87-115. In press.
Vogel, D.G., Rabinovitch, P.S., and Mottet, N.K. 1986. Methyl-mercury effects on cell cycle
kinetics. Cell and Tissue Kinetics 19: 227-242.
Vogelbein, W.K., Fournic, J.W., van Veld, P.A., and Huggett, RJ. 1990. Hepatic neoplasms in
the mummichog (Fundulus heteroclitus) from a creosote-contaminated site. Cancer
Research 50: 5978-5986.
Vollenwieder, RA. 1976. Advances in defining critical loading levels for phosphorus in lake
eutrophication. Memorie Institute Italiano di Idrobiologia 33: 53-83.
Vollenwieder, RA. 1979. Das Nahrstoffbelastungskonzept als Grundlage fur den externen
Eingriff in den Eutrophierungsprozess stehender Gewasser und Talsperren. Z. Wasser-u.
Abwasser-Forschung 12: 46-56.
Voogt, P.A., den Besten, P.J., Kusters, G.C.M., and Messing, M.W.J. 1987. Effects of cadmium
and zinc on steroid metabolism and steroid level in the sea star (Asterias rubens L.~)
Comprehensive Biochemistry and Physiology 84B: 83-89.
Voorman, R. and Aust, S.D. 1987. Specific binding of polyhalogenated aromatic hydrocarbon
inducers of cytochrome P-450d to the cytochrome and inhibition of its estradiol 2-
hydroxylase activity. Toxicology and Applied Pharmacology 90: 69-78.
Voorman, R. and Aust, S.D. 1989. TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin) is a tight
binding inhibitor of cytochrome P-450d. Journal of Biochemical Toxicology 4: 105-9.
197
-------
Vos, J.G. and de Roy, T.H. 1972. Immunosuppressive activity of a polychlorinated biphenyl
preparation on the humoral immune response in guinea pigs. Toxicology and Applied
Pharmacology 21: 549-555.
Wakeling, A.E. and Visek, W J. 1973. Insecticide inhibition of 5a-dihydrotestosterone binding
in the rat ventral prostate. Science 181: 659-661.
Wariishi, M., Suzuki, Y., and Nishiyama, K. 1986. Chlordane residues in normal human blood.
Bulletin Environmental Contamination and Toxicology 36(5): 635-643.
Warngard, L., Fransson, R., Drakenberg, T., Flodstrom, S., and Ahlborg, U. 1988. Calmodulin
involvement in TPA and DDT induced inhibition of intercellular communication.
Chemistry and Biological Interactions 65: 41-49.
Warriner, J.E., Mathews, E.S., and Weeks, BA. 1988. Preliminary investigations of the
chemiluminescent response in normal and pollutant-exposed fish. Marine Environmental
Research 24: 281-284.
Weber, L.W., Greim, J., Rozman, K.K. 1987. Metabolism and distribution of [14C]glucose in rats
treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Journal of Toxicology and
Environmental Health 22(2): 195-206.
Weber, L.W., Lebofsky, M., Stahl, B.U., Gorski, J.R., Muzi, G., and Rozman, K. 1991. Reduced
activities of key enzymes of gluconeogenesis as possible cause of acute toxicity of
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in rats. Toxicology 66: 133-144.
Weeks, BA., and Warriner, I.E. 1984. Effects of toxic chemicals on macrophage phagocytosis
in two estuarine fishes. Marine Environmental Research 14: 327-335.
Weeks, B.A., Warriner, I.E., Mason, P.L., and McGinnis, D.S. 1986. Influence of toxic chemcials
on the chromatactic response of fish macrophages. Journal of Fish Biology 28: 653-658.
Weimeyer, S., Lamont, T., Bunck, S., Sindelar, C, Gramlich, F., Fraser, J., and Byrd, M. 1984.
Organochlorine pesticide, polychlorobiphenyl, and mercury residues in bald eagle eggs-
1969-1979- and their relationship to shell thinning and reproduction. Archives of
Environmental Contamination and Toxicology 13: 529-549.
Welch, R.M., Levin, W., and Conney, A.H. 1969. Estrogenic action of DDT and its analogs.
Toxicology and Applied Pharmacology 14: 358-367.
Wells, M.L., Mayer, L.M., and Guillard, R.R.L. 1991. Evaluation of iron as a triggering factor
for red tide blooms. Marine Ecology Progress Series 69: 93-102.
198
-------
Weseloh, D., Teeple, S., and Gilbertson, M. 1983. Double-crested cormorants of the Great
Lakes: egg-laying parameters, reproductive failure, and contaminant residues in eggs,
Lake Huron 1972-1973. Canadian Journal of Zoology 61: 427-436.
Wetzel, R.G. 1983. Limnology. Philadelphia, PA: Saunders.
Wickizer, T, Brilliant, L., Copeland, R., and Tilden, R. 1981. Polychlorinated biphenyl
contamination of nursing mothers1 milk in Michigan. American Journal of Public Health
71: 132-137.
Wigfield, D.C. and Eatock, S.A. 1992. The effect of metals on the activity of L-phenylalanine
hydroxylase. Journal of Trace Elements and Electrolytes in Health and Disease 4: 143-
146.
Winek, C, Fochtman, F., Bricker J., Wecht, C.H. 1981. Fatal mercuric chloride ingestion.
Clinical Toxicology 18: 261-266.
Winneke, G., Brockhaus, A., Collet, W., and Kramer, U. 1989. Modulation of lead-induced
performance deficit in children by varying signal rate in a serial choice reaction task.
Neurotoxicology and Teratology 11(6): 587-592.
Wisconsin Department of Health. 1987. Wisconsin Division of Health and State Laboratory of
Hygiene. Study of sport fishing and fish consumption habits and body burden levels of
PCBs, DDE, and mercury of Wisconsin anglers.
Wisconsin Sea Grant Program. 1976. ABCs of PCBs. Public Information Report #WIS-SG-76-
125, University of Wisconsin. Madison, WI.
Wolfe, DA., Monhahan, R., Stacey, P.E., Farrow, D.R.G., and Robertson, A. 1991.
Environmental quality of Long Island Sound: assessment and management issues.
Estuaries 14: 224-236.
Wong, K.C. and Hwang, M.Y. 1981. Children bom to PCB poisoned mothers. Clinical Medicine
(Taipai) 7: 83-87.
Woodley, T.H., Brown, M.W., Kraus, S.D., and Gaskin, D.E. 1991. Organochlorine levels in
North Atlantic right whales (Eubalena glacialis) blubber. Archives of Environment and
Contamination Toxicology 21: 141-145.
World Health Organization (WHO). 1976. Environmental Health Criteria 1: Mercury.
World Health Organization (WHO). 1984. Environmental Health Criteria 44: Mirex.
199
-------
World Health Organization (WHO). 1989. Environmental Health Criteria 88: Polychlorinated
Dibenzo-Para-Dioxins and Dibenzofurans. Geneva.
World Health Organization (WHO). 1990. Environmental Health Criteria 101: Methylmercury.
Geneva.
World Health Organization (WHO). 1991. Environmental Health Criteria 118: Inorganic Mercury.
Geneva.
Wren, CD., Hunter, D.B., Leatherland, J.F., and Stokes, P.M. 1987. The effects of
polychlorinated biphenyls and methylmercury, singly and in combination on mink II:
reproduction and kit development. Archives of Environmental Contamination and
Toxicology 16: 449-454.
Wright, DA., Hartwell, S.I., and Savitz, J.D. 1992. Low-level effects of toxic chemicals on
Chesapeake Bay organisms. Pp. 45-74 in: Perspectives on Chesapeake Bay, 1992:
Advances in Estuarine Sciences. Scientific and Technical Advisory Program. Chesapeake
Bay Program. Publication No. 143.
Wulff, F., Stigebrandt, A., and Rahm, L. 1990. Nutrient dynamics of the Baltic Sea. Ambio 19:
126-133.
Yoshida, M., Satch, H., Kishimoto, T., and Yamamura, Y. 1992. Exposure to mercury via breast
milk in suckling offspring of maternal guinea pigs exposed to mercury vapor after
parturition. Journal of Toxicology and Environmental Health 35: 135-139.
Zacharewski, T., Harris, M., and Safe, S. 1991. Evidence for the mechanism of action of the
2,3,7,8-tetrachlorodibenzo-p-dioxin-mediated decrease of nuclear estrogen receptor
levels in wild-type and mutant mouse Hepa Iclc7 cells. Biochemical Pharmacology
41(12): 1931-1939.
Zacharewski, T., Harris, M., Biege, L., Morrison, V. Merchant, M., and Safe, S. 1992. 6-
Methyl-l,3,8-trichlorodibenzofuran (MCDF) as an antiestrogen in human and rodent
cancer cell lines: evidence for the role of the Ah receptor. Toxicology and Applied
Pharmacology 113(2): 311-318.
Zeilmaker, M. and Yamasaki, H. 1986. Inhibition of functional intercellular communication as
a possible short-term test to detect tumor-promoting agents: results with nine chemicals
tested by dye transfer assay in Chinese hamster V79 cells. Cancer Research 46(121):
6180-6186.
Zell, M. and Ballschmiter, K. 1980a. Baseline studies of the global pollution: n. Global
occurrence of hexachlorobenzene (HCB) and polychlorocamphenes (toxaphene) (PCB) in
biological Samples. Fresenius Zeitung Analitische Chemie 300: 387-402.
200
-------
Zell, M. and Ballschmiter, K. 1980b. Baseline studies of the global pollution: m. Trace analysis
of polychlorinated biphenyls (PCB) by EDC glass capillary gas chromatography in
environmental samples of different trophic levels. Fresenius Zeitung Analitische Chemie
304: 337-349.
Zhong-Xiang, L., Kavanagh, T., Trosko, J., and Chang, C. 1986. Inhibition of functional
intercellular communication in human teratocarcinoma cells by organochlorine pesticides.
Toxicology and Applied Pharmacology 83: 10-19.
201
------- |