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A TOXICITY INDEX FOR ASSESSING
THE SUITABILITY OF STREAMS FOR AQUATIC LIFE
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14 July 1988
by
Richard E. Sparks
Aquatic Biology Section
Illinois Natural History Survey
River Research Laboratory
Box 599
Havana, Illinois 62644
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Toxicity index July 14, 1988 Page 1
A TOXICITY INDEX FOR ASSESSING THE SUITABILITY OF STREAMS FOR
AQUATIC LIFE
R.E. Sparks
INTRODUCTION
Habitat quality, water quality, and biotic interactions all affect
aquatic organisms. Streams have been classified according to the suitability
of the water or habitat for aquatic life (water depth, flow velocity,
substrate, temperature, concentration of oxygen and pollutants) or according
to characteristics of the indigenous populations (species diversity, presence
or absence of indicator species or guilds). Although some aquatic ecologists
consider water quality a part of habitat quality, it is useful in this paper
to make a distinction between chemical characteristics of the water which are
measured routinely in water quality monitoring programs and the physical
characteristics of the habitat which are not. While population data can
indicate that a problem exists, the cause of the problem is not always easily
determined from available information on water quality or habitat quality. If
the causes are unknown, it is difficult to design measures to restore the
biological quality of the stream. The biological significance of water
quality data is often obscure, particularly because factors are usually
considered one at a time (does the concentration of factor X exceed the water
quality standard for aquatic life?), whereas organisms are exposed to many
factors simultaneously. Water chemistry is usually measured in grab samples
taken once a month, or even less frequently, so it usually is impossible to
determine how long organisms were exposed to stressful factors. The degree of
stress depends not only on concentration but also on duration of exposure.
The toxicity index described in this paper sums component toxicities
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contributed by 20 common pollutants. The algorithms account for known effects
of environmental factors which modify toxicity (temperature, dissolved oxygen,
hardness or alkalinity). The toxicity index can be coupled to hydrologic and
water quality models to estimate exposure durations as well as toxicity
magnitudes, and to develop empirical relationships between index values,
exposure durations, and fish community structure.
Although toxicity indices have been developed in both North America and
Europe, research in Illinois has contributed substantially to verification,
refinement and novel application of toxicity indices, including application to
stream classification.
This paper describes: (1) the background, assumptions, and limitations of
the toxicity index, (2) an example of the computation of the toxicity
contributed by one component (ammonia), (3) application of the toxicity index
to two rivers in Illinois, for the purposes of classifying reaches according
to suitability for fish, determining which factors contribute the most
toxicity, and for empirical determination of the relationship between the
toxicity index and characteristics of the fish populations, and (4) future
applications and improvements to the index.
ASSUMPTIONS AND LIMITATIONS
The Additive Assumption
Sprague and Ramsay (1965), Lloyd (1965), and Sprague (1970) were the
first to propose the toxic units approach to predicting the joint toxicity of
mixtures of common industrial and municipal pollutants. They expressed the
separate toxicant concentrations as fractions of their lethal threshold
concentrations and assumed the joint lethal effects were additive. The units
were referenced to the species used to determine the lethal thresholds in
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Digitized by the Internet Archive
in 2010 with funding from
CARLI: Consortium of Academic and Research Libraries in Illinois
http://www.archive.org/details/toxicityindexforOOspar
Toxicity index July 14, 1988 Page 3
laboratory bioassays. The bluegill is the reference organism in the Illinois
toxicity index because it is widespread in Illinois (it is the state fish) and
is commonly used in bioassays (Lubinski 1981; Lubinski and Sparks 1981;
Lubinski et al. 1974; Muchmore et al. 1979; and Brigham and Hey 1981). If the
concentration of zinc in water which kills 50% of the exposed population of
bluegill sunfish (Lepomis macrochirus) in 96 hours is 8 mg/1 (the 96-hour
LC50), then 8 mg/1 is considered to be 1.0 toxic unit, or more specifically,
1.0 bluegill toxicity unit .BGTU (Lubinski et al. 1974). A zinc concentration
of 4 mg/1 in a mixture therefore is 0.5 BGTU, and is the component toxicity
attributable to zinc. If a stream contained 0.5 BGTU of zinc and 0.5 BGTU of
copper, the water in this stream has a toxicity index value of 1.0 (0.5 + 0.5
= 1.0) and is predicted to be lethal to fish in 96 hours of exposure.
Two controversies developed over this approach. One was whether
combinations of common pollutants were in fact additive, more than additive,
or less than additive in contributing to lethality. The other was whether the
toxicity index could be empirically related to sublethal effects on fish
populations, e.g., failure of reproduction or changes in species occurrence or
dominance.
The net results of many laboratory and field tests of the additive
assumption are that very toxic mixtures (1.0 toxic unit) are more-than-
additive, as measured by survival time; mixtures where component toxicities
exceed 0.2 generally are additive, as measured by lethal thresholds; mixtures
where component toxicities are less than 0.2 usually fail to add because the
components apparently contribute no acute toxicity to the mixture; and
reproducing populations of native fishes can be maintained where toxicity
indices are less than 0.2, and habitat (cover, substrate, water depth and
velocity) and biotic factors (food supply, predators, competitors, parasites
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and disease organisms) are not limiting. These conclusions are supported by
literature which will be reviewed briefly here, because of the central
importance of the additive assumption to the toxicity index, and by the
results of the two applications which are described later.
The archetypical example of a more-than-additive toxic effect was
furnished by Doudoroff in 1952. He demonstrated that Pimephales sp. could
withstand 8.0 ppm of zinc alone or 0.2 ppm of copper alone for 8 hours, but
the combination of only 1.0 ppm zinc with 0.025 ppm copper killed most of the
fish in 8 hours. This example has been cited frequently and used as a warning
of the type of more-than-additive effects to be expected if more and more
pollutants are introduced to the environment, without an assessment of their
joint effects (Cairns 1957). More-than-additive effects of zinc-copper
mixtures have also been demonstrated by other investigators (Lloyd 1961b).
It is therefore surprising that Herbert stated in 1965 that no cases had
been found in which the toxicity of mixtures of poisons commonly found in
sewage and industrial wastes (copper, zinc, lead, phenol, ammonia, and
cyanide) were appreciably more than additive. This apparent contradiction is
resolved by looking more closely at the concentrations and exposure times used
in the experiments. As Sprague and Ramsay (1965) pointed out, survival times
in mixtures where the total toxicity appreciably exceeds 1.0 toxic unit
(Doudoroff's example) are shorter than expected. When lethal thresholds (96-
hr LC50's or LT50's) are measured, the toxicities of mixtures of heavy metals
and other toxicants generally do add up in laboratory experiments. However,
there are exceptions as shown below.
Sprague (1970) pointed out that the toxicity of mixtures of phenol,
ammonia, and zinc was overestimated by Brown et al. (1969) in three out of
four cases where two of the toxicants were present at 0.10 to 0.14 toxic
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units. When each toxicant was present at 0.2 toxic units or more, the
toxicities did add up. These data suggest that the threshold for acute toxic
effect postulated by Lloyd (1965) does exist, at least for metals, and that it
may be approximately 0.2 toxic unit.
Results using mixtures other than metals alone are ambiguous. In
Illinois Sparks and Anderson (1977) reported that a toxicity index
underestimated the lethality of a mixture of linear alkyl sulfonate (LAS, a
detergent), ammonia, and zinc by approximately 50%. In contrast, Esvelt et
al. (1971) accurately estimated the toxicity of wastes entering San Francisco
Bay by adding the toxic contributions of methylene blue-active substances
(MBAS, mostly detergents) and ammonia. Herbert (1962) noted an 82.5 percent
agreement between predicted and observed toxicities in field tests, although
Sprague (1970) noted that the limit set for agreement was rather wide. In
fresh-water reaches of four rivers in England, actual 48-hr LC50s approximated
65% of predicted values, and in the saline estuaries, prediction further
underestimated actual toxicity (Brown et al. 1970). Lloyd and Jordan (1964)
found their index consistently underestimated the toxicity of sewage effluents
and that the relation between the predicted and observed toxicity was
described by the function:
y = 1.25x - 0.59
where y is the observed toxicity and x the predicted toxicity.
Some of the underestimates of toxicity reported in the older literature
probably result from bioassays which were run for arbitrary time periods,
e.g., 8-, 24-, or 48-hours, which were too brief for full uptake of the
toxicant and full exertion of the toxic effect in the test populations,
particularly at the lower concentrations. The individuals in every test
population differ in tolerance to the toxicant, and the difference is
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expressed in survival time. The thresholds estimated from short-term
bioassays could underestimate actual thresholds. Modern practice is to
continue an acute bioassay for at least 96 hours, or better, until all
mortality (or other toxic effect) has ceased for at least 24 hours.
Even with these limitations of the additive assumption however, most of
the predicted toxicities of mixtures differed from the actual toxicities by no
more than 50%. Considering that the modes of action of many of the toxicants
are unknown, let alone the modes of interaction with other toxicants and with
modifying factors in the water or in the organisms, it is remarkable that the
error is not higher. It is likely that the error will be reduced as more is
learned about interactions, which then can be incorporated in the toxicity
index. Acceptance of a 50% error appears preferable to the alternative of not
using the existing bioassay literature and chemical monitoring data to
estimate the joint effects of toxicants in lakes and streams. In cases such
as spills of complex wastes, where the predicted toxicity is several times
greater than 1.0, a 50% error is insignificant. At the other extreme, field
evidence from a variety of polluted rivers in England and Illinois supports a
relatively narrow range of 0.2-0.4 for a threshold below which populations of
native fishes can sustain themselves, if other factors are not limiting (this
paper; Brigham and Hey 1981; Herbert et al. 1965; and Brown 1970).
Bioaccumulation
By definition, bluegill toxicity indices sum 96-hour LC50 values and thus
pertain only to acute toxicity, or to empirically determined relationships
between index values and fish populations in streams. However, similar
indices can be based on chronic effects when sufficient data are developed.
The index does not assess effects of toxicants which accumulate in organisms
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from levels in water which are too low to have acute effects. The index
therefore does not identify chemicals which might not affect fish but might
affect consumers of fish.
Choice of Species and Life Stage
The present toxicity index is based on juvenile and adult life stages of
the bluegill because most of the available toxicological information was for
these stages. Additional toxicity indices, however, could be based on
sensitive embryonic and larval stages of organisms (e.g. Reinbold and
Pescitelli 1982) representing a variety of trophic levels or functional groups
(e.g. Anderson et al. 1978) within aquatic systems. Multiple indices would be
better predictors of species replacements or ecosystem-level effects.
Limitations of the Data Bases
Algorithms for computing component toxicities can be continually updated
as more toxicity data become available on newly introduced chemicals and on
interactions and modifying factors, but in the meantime, there are data gaps.
For example, I could locate no information on the toxicity of fluoride to
bluegills. Neuhold and Sigler (1960), however, reported 96-hr LC50s for
fluoride-sensitive rainbow trout and fluoride-tolerant common carp, and it was
reasonable to assume bluegill would be intermediate in sensitivity.
Automated samplers, event-triggered sampling (during a flood, drought, or
spill), and water quality models (calibrated with available data) can overcome
the limitations imposed by the relatively infrequent sampling (usually once a
month, at best) characteristic of most water quality monitoring programs. It
is particularly important to know the duration of exposure because fish can
survive brief exposures to conditions (low oxygen, lethal concentration of a
toxicant) which eventually would kill them. Another limitation of the water
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quality data base is that total concentrations are measured, rather than toxic
fractions. Doudoroff et al. (1966) demonstrated that molecular cyanide (HCN)
is the toxic agent in solutions containing ionized cyanide (CN~) and cyanide-
metal complexes, but only total cyanide is measured by IEPA and it may not be
reliable to compute the molecular fraction of complex cyanide solutions in
streams, even when the general chemical composition and pH are known
(Doudoroff 1976). If nontoxic cyanide complexes are present, use of the total
cyanide concentration overestimates the toxicity. These limitations could be
overcome with better analytical techniques or more sophisticated programs for
calculating chemical equilibria in complex mixtures.
CALCULATION OF COMPONENT TOXICITIES: AN EXAMPLE USING AMMONIA
For some toxicants, the effects of environmental factors on the chemical
form and concentration of the toxicant, and on the resistance of the aquatic
organism, have been determined in the laboratory and can be accounted for in
the toxicity index. An example, using ammonia, is described next.
Ammonia is a common pollutant in Illinois waters. It is formed by the
breakdown of urea, and hence is found in effluents from livestock confinement
areas and sewage treatment plants. Ammonia, in several chemical forms, is
stored, transported by pipe, truck, rail and barge, and applied to
agricultural lands as a nitrogen source for crops. It occurs in effluents
from refineries and munitions industries.
Effects of Modifying Factors on Chemical Equilibria
Ammonia in water exists in a toxic, un-ionized form, NHo, and a non-toxic
ionized form, NH^ + , with the equilibrium between the two forms governed by pH,
temperature, and salinity, although in freshwater the salinity effect can be
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ignored (Emerson et al. 1975). The total ammonia-nitrogen concentration
(ionized + un-ionized) is measured in most water quality monitoring networks,
but the proportion which exists in the toxic state can be determined from the
field pH and temperature, using the tables or two equations provided by
Emerson et al. (1975: 2382).
NHj-N = total ammonia-N x 1
1 + antilog (pka -pH)
where pka = the negative log of the ionization constant:
pka = -0.03229 (temp °C) + 10.05333
Effects of Modifying Factors on Sensitivity of the Fish
The toxicity of un-ionized ammonia to fish varies with the size of the
fish and the temperature and dissolved oxygen concentration of the water
(Roseboom and Richey 1977; Reinbold and Pescitelli 1981; Merkens and Downing
1957). The toxicity of ammonia increases at low temperatures probably because
the overall metabolic rate of the fish is lower, and their ability to excrete
ammonia in their urine is reduced. Fish pick up ammonia from the water via
their gills, and form ammonia within their tissues as a waste product of
protein metabolism. When the rate of ammonia uptake and production exceeds
the rate of ammonia excretion, the internal concentration of ammonia rises to
lethal levels (Fromm 1970; Brockway 1950). Small fish are more sensitive to
ammonia than large fish, presumably because the ratio of gill surface
available for ammonia uptake to body volume or mass is greater in the smaller
fish. The toxicity of every chemical used in the toxicity index, including
ammonia, increases at low dissolved oxygen concentrations, presumably because
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low oxygen itself is a stressor, and virtually any reduction below saturation
reduces the metabolic scope for excretion or detoxification. As ambient
oxygen concentrations decline, some fishes can compensate to some extent by
increased ventilation frequency or volume (Marvin and Heath 1968), but
probably at the expense of other metabolic activities, including ammonia
excretion.
The 96-hour LC50 for un-ionized ammonia (as NH^-N) was regressed against
fish weight (fwt) and water temperature (temp), using Illinois data of
Reinbold and Pescitelli (1981) and Roseboom and Richey (1977). The lowest and
highest temperatures used by these investigators were 4 and 28°C, and the
regression equation is not extrapolated beyond the range of these data:
if temp < 4 °C, then the 96-hour LC50 = i (° 26639 fwt + -025353(4) - .67645)
if temp > 28 °C, then the 96-hour LC50 = i (- 026639 fwt + -025353(28) - .67645)
otherwise, the 96-hour LC50 = i (- 026639 fwt + •025353(temp) - .67645)
Next, the LC50 is adjusted to reflect the increased toxicity of un-
ionized ammonia at dissolved oxygen (DO) concentrations below saturation
(Merkens and Downing 1957):
LC50 = LC50 (at 100% saturation) x .013297 (DO, % saturation) - .32965
Unfortunately, only 2 levels of dissolved oxygen were tested. I assumed a
linear relationship between the LC50 and DO.
Application to an Ammonia Spill in the Illinois River
For the purposes of this example, fish weight is set at 1 gram, which
would probably be at the lower end of the average weight for bluegills in
their first year of life in January (Carlander 1977) when 622,000 gallons of
urea ammonium nitrate solution spilled into the Illinois River at Seneca,
Illinois. A dispersion model indicated that the total ammonia concentration
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6.5 river miles downstream, at Marseilles, was 47 mg/1 for approximately 19
hours (personal communication, 8 April 1988, Mr. Thomas Butts, Professional
Scientist and Assistant Head, Water Quality Section, Illinois State Water
Survey, Peoria, Illinois).
Sensitivity of Ammonia Component Toxicity to Modifying Factors.
Dissolved oxygen levels have a marked effect on the predicted toxicity of 47
mg/1 total ammonia-N (Table 1). At saturation, the ammonia spill exceeded
twice the lethal concentration, but at 13% saturation, the component toxicity
was greater by 3 orders of magnitude, and would probably have killed fish
within a few hours at summer temperatures. At colder temperatures, less of
the total ammonia exists in the toxic, un-ionized form, but the sensitivity of
the fish increases, so the net change in component toxicity is rather small,
from 2.69 to 1.90, as the temperature declines from 28 ° C to 4 ° C (Table 1).
At cold temperatures, fish may be exposed to ammonia longer than it takes a
spill to pass a fixed point, because they lose their equilibrium and float
upside down (Reinbold and Pescitelli 1981) and thus would be carried along in
the spill.
Conclusions
Even assuming the DO was at saturation, and no other toxicants were
present, the toxicity attributable to the spill was approximately twice the
lethal threshold at the cold temperatures expected in January. Fish which
lost their equilibrium were exposed to these lethal concentrations longer than
19 hours, and very probably died.
The component toxicity approach was a rational way of integrating
information on environmental factors, chemical concentrations, and
susceptibility of organisms to assess the probable degree of harm to aquatic
life caused by this ammonia spill. The next application examines toxicities
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contributed by 20 components in the Illinois River and the connecting Chicago-
area canals, collectively known as the Illinois Waterway.
APPLICATION TO THE ILLINOIS WATERWAY
The Illinois Waterway extends from the highly altered rivers and
waterways in Chicago downstream 326 miles via the Illinois River to the
Mississippi River upstream of St. Louis. The numbers and kinds of native
gamefish and their condition declines in the upstream direction, towards
Chicago, where fish populations are dominated by a few species, including the
introduced carp and goldfish (Sparks and Starrett 1975).
Procedures
In the years 1972-1974, the toxicity index was applied to the Illinois
Waterway, to determine whether high toxicity values were associated with the
degraded fish populations in the upstream reaches, and to determine what
components contributed the most toxicity. Field tests, using bluegills
exposed to river water for 4 days in cages in the river and in aerated plastic
pools on shore (where the water was renewed by pumping), were conducted at an
upstream (Dresden) and a downstream (Beardstown) site in 1974 (Lubinski et al.
1974). Controls were maintained nearby in cages in less polluted tributaries.
Water quality data were obtained from the Illinois Environmental
Protection Agency (IEPA) for samples taken 4 to 13 times per year at 20
stations along the Illinois Waterway. The same data were obtained by the
Natural History Survey on samples taken at least once daily during the field
tests.
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Results
Field bioassays. The field bioassays confirmed that no short-term
mortalities attributable to toxicity occurred in the two reaches of the
Illinois River or in the tributaries where the daily index values were less
than 0.14 (Lubinski et al. 1974). No mortalities occurred at Dresden. At the
Beardstown site, at the mouth of the Sangamon River, there were no mortalities
in the aerated swimming pool receiving Illinois River water, but 22% of the 50
test fish confined in the cage in the Illinois River died, probably because of
the stress of swimming against a relatively high current velocity (0.4-0.6
m/sec) or being forced against the mesh by the current (Lubinski et al. 1974).
Twelve percent of the fish died in a backwater area receiving flow from the
Sangamon River where the water levels were falling and the cage had to be
lifted from the mud and moved to deeper water several times, stressing the
fish.
Mean Toxicity Indices. The mean annual toxicity indices in the Illinois
Waterway (the sum of the toxicity indices at each station, divided by the
number of samples taken that year) generally were below 0.1 at the station
farthest upstream, which receives clean water from Lake Michigan via a
navigation lock, well above 0.2 through the Chicago waterways, and 0.1 or less
starting where the Chicago Sanitary and Ship Canal joins the Des Plaines River
(Figure 1). The year 1972 was an exception, with elevated toxicity indices at
Peoria.
Maximum Toxicity Indices. Extreme events inevitably go undetected
because of the small number of water quality samples taken per year by the U.
S. Geological Survey and the Illinois Environmental Protection Agency (12 or
less). If it takes approximately 5 minutes to fill the sample bottle each
month, there are 43,195 minutes in which no samples were taken (60 minutes x
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24 hours x 30 days = 43,200, less 5 minutes = 43,195). The sample represents
conditions occurring during only 1 / 10,000th of the month. Maximum annual
toxicity indices partially compensate for this limitation by adding the peak
toxicities contributed by each component during the year. The assumption is
that the highest component toxicities occurred simultaneously, and the maximum
index approximates "worst case" conditions. Such an assumption is not
completely unreasonable because: (1) if one toxicant is at high concentration
because of an excessive discharge, several others usually are too, because
industrial wastes typically are complex mixtures of pollutants, and (2) during
low flow conditions, all wastes are more concentrated because there is less
dilution capacity. (A better, but more costly simulation procedure, in terms
of data and programming requirements, which does not make the "simultaneous
maxima" assumption, is described in application 2 below.) Maximum toxicity
indices approach 0.5 in the upper Illinois Waterway (except for the uppermost
station, near Lake Michigan), and generally decline downriver, with the
exception of the Peoria stations in 1972 (Figure 1).
Component Toxicity: Cyanide. The 1972 peak in toxicity at Peoria is
explained largely by the cyanide component (Figure 2), with some contribution
from zinc and copper (relatively high values, compared to other years, did
occur together in one sample in this case). The spiky, highly variable
cyanide pattern is typical of toxicants which are accidentally or sporadically
introduced, and it is fortuitous that the monthly sampling in 1972 happened to
detect an apparent spill which was moving downstream, and which probably
originated from an industrial source in Peoria.
Component Toxicity: Ammonia. The dominant contributor to the pattern of
high toxicity in the Chicago waterways and declining toxicity downstream is
ammonia (Figure 2), which is continuously released from the sewage treatment
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plants in Chicago. Ammonia probably declines downstream because of dilution
by tributaries, uptake by aquatic plants in the upper Illinois Waterway, and
conversion to nitrate by bacteria in the water. Toxicity in portions of the
Illinois Waterway in Chicago could be reduced below an index of 0.2 by
increased diversion of clean dilution water from Lake Michigan, but the
diversion increase would have some potentially deleterious effects, including
possible scouring of toxicant-laden sediments from the waterways downstream
into more biologically productive areas (Havera et al. 1980).
Despite improvements in sewage treatment in the Chicago River since the
early 1970s (Macaitis et al. 1987), ammonia remains a problem in much of the
Illinois River. Unpublished data from the U. S. Fish and Wildlife Service
(personal communication, 29 February 1988, Richard Ruelle, Aquatic
Toxicologist, Environmental Services Section, U. S. Fish and Wildlife Service
Field Office, Rock Island, Illinois) indicates that sediments in backwaters at
least as far as 200 miles downstream from Chicago contain ammonia which is
released in sufficient quantities when the sediments are agitated to kill fish
in 96-hour laboratory bioassays. Sediments are frequently agitated by wind-
and boat-generated waves (Jackson and Starrett 1959; Sparks and Starrett
1975). Sparks (1984) reported that un-ionized ammonia concentrations in
excess of the lethal level for fingernail clams (0.06 mg/1 NH^-N) occurred in
the Illinois River at Marseilles, Hennepin and Lacon in 1980, probably because
uptake of CC»2 and HCO3- during algal blooms raised the pH above 8 and
increased the proportion of ammonia existing in the toxic un-ionized form.
Fingernail clams are an important food for diving ducks and bottom-feeding
fish, both of which were adversely affected by the die-off of the clams and
other mud-burrowing invertebrates in a 100-mile reach of the Illinois River in
1958 and by their continuing failure to recolonize (Mills et al. 1966; Sparks
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1984).
Summary of Application 1
In summary, the toxicity index proved useful in interpreting the water
quality data collected by the IEPA on the Illinois Waterway to determine what
toxicants are acutely limiting to fish and should be controlled to restore
fish populations. Index values above 0.2 in the upper Illinois Waterway are
associated with depauperate fish populations in generally poor body condition.
The toxicity index falls below 0.2 downstream, where more species of native
fishes, in better condition, occur. The principal contributor to the total
acute toxicity is ammonia, with sporadic contributions from cyanide. Ammonia
originates from such a widely dispersed, large capacity sewage system
(Chicago) that pulses are absorbed and ammonia loading is relatively constant,
but downstream variations in temperature, pH, and dissolved oxygen alter the
component and overall toxicities. Therefore, an important limitation was the
infrequency of water quality sampling (only 4 to 13 times per year), which
made it difficult to detect extreme conditions or determine how long fish were
exposed. The next application describes how water quality modeling can be
used to overcome this limitation and, when coupled with the toxicity index,
used to predict the effects of alternative pollution control measures on fish
populations.
APPLICATION 2: EVALUATING STRATEGIES FOR WATER QUALITY MANAGEMENT
Objectives
The objectives of this project were (1) to explore the relationship
between fish communities and their physical and chemical environments, using
the good water quality and fish population data sets and a calibrated water
quality model available for the DuPage River in northeastern Illinois, (2) to
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develop a continuous toxicity function for relating the status of these
communities to spatial and temporal patterns of physical and chemical events,
and (3) to use the resulting toxicity functions to predict effects of
alternative pollution control measures on fish populations (Brigham and Hey
1981).
Procedure
Three stream reaches of the DuPage River were selected for modeling and
analysis, based on known differences in fish faunas. One reach supported a
mixed community of 17 species, including bluegill and carp, (hereafter
referred to as the bluegill reach), one supported only 6 species, excluding
bluegill and dominated by carp (carp reach), and the third reach was fishless
(Brigham and Hey 1981). The toxicity index was calculated for each reach at
1-hour intervals over a simulated span of 3 years, using water quality values
generated by hydrologic and water quality models which were calibrated for the
reaches. The models were developed by Hydrocomp, Inc., Palo Alto, California
and implemented by the Northeastern Illinois Planning Commission. The
toxicants were un-ionized ammonia, cyanide, lead, zinc, copper, linear
alkyl sulfonate detergents (LAS), and total residual chlorine (TRC). After
initially observing the magnitude and variability of the toxicity function
generated for each reach (Figure 3), certain water quality input variables
were modified to simulate different management practices on each stream reach.
Results
Critical Thresholds. The toxicity indices for the three stream reaches
on 17 March 1972 are typical of the run simulating the period from 1 October
1970 to 30 September 1973 (Figure 3). Toxicity levels regularly exhibited in
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these reaches compared well with the results of previous investigators, who
indicated that the effects of acute toxicity in altering the species
composition of fish communities became measurable in streams at levels of 0.2
to 0.4 toxic units (Lloyd and Jordan 1964; Edwards and Brown 1966).
Relationships between Frequency, Duration and Intensity of Exposure and
Fish Populations. Perhaps more important than the typical values however, are
the frequency and duration of episodes where toxicities exceeded the lethal
value of 1.0 and the empirically determined threshold of 0.2-0.4 (Table 2). In
the fishless reach , the toxicity index made 194 excursions above 3.0 lasting 1
hour or more during the 3 years, 39 excursions lasting 24 hours or more, and
21 lasting 96 hours or more. The carp reach exhibited toxicity levels of 1.0
unit for 96 hours or more on 14 occasions, whereas the bluegill reach never
exceeded 0.3 for even 1 hour. There were only 2 occasions during the
simulated 3-year period when the toxicity index in the bluegill reach was
between 0.25 and 0.30 for 24 hours or more. The average length of these 2
excursions was 36.5 hours (Brigham and Hey 1981).
Simulation of Management Alternatives
Assume a management goal of changing the fishless reach to a reach
capable of supporting fish. A typical management plan might include: (1)
reduction of ammonia concentrations from wastewater treatment plants to 1.5
mg/1 during the summer and 4.0 mg/1 during winter, (2) elimination of combined
(stormwater and sewage) sewer overflows, (3) reduction of sediment oxygen
demand, and (4) moderate increase of dissolved oxygen in the wastewater
effluents or in the stream (Brigham and Hey 1981).
This plan primarily targets ammonia toxicity, which is reduced by 3
orders of magnitude, from 20.2 to 0.023 (Table 3). The mean toxicity index
declines by an order of magnitude, from 23.0 to 2.12 (Table 3), but still
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significantly exceeds the mean index of 1.04 in the carp reach (Brigham and
Hey 1981). The largest remaining contributor to the total toxicity is
chlorine (mean component toxicity = 1.94). Cessation of effluent chlorination
would reduce the mean index to 0.186, which is very close to the mean of 0.115
in the bluegill reach. If no excursions above 0.3 occurred, and no factors
other than toxicity are limiting, a mixed community of native fishes probably
could be maintained.
This simulation was run in 1981, and the Illinois Environmental
Protection Agency has since abandoned the requirement for effluent
chlorination, based on evidence that there would be little additional public
health risk from infectious diseases and much improvement in water quality for
aquatic life. The latter evidence included studies which employed or referred
to the toxicity index (Muchmore et al. 1979; Dreher 1981; Hey, Pappas and Cox
1980; and Hey et al. 1982). Fish populations in the Chicago waterways have
shown recent improvement following discontinuation of effluent chlorination
(personal communication, 1 March 1988, Mr. Samuel Dennison, Fisheries
Biologist, Metropolitan Sanitary District of Greater Chicago).
FUTURE DEVELOPMENT
The toxicological data base for the index was last revised in March 1981.
The programs for computing toxicity indices from IEPA water quality data are
written in BASIC for a Tektronix 4051 microcomputer and a CYBER 175 at the
University of Illinois (Lubinski 1981). The University is replacing the CYBER
and the Tektronix 4051s are no longer in common use. The algorithms should be
rewritten, using updated data, for IBM- or Apple-compatible personal
computers, and a new user's guide prepared.
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The revised index then should be applied to reaches intermediate in
toxicity between the bluegill reach and the carp reach of the DuPage River
study (Brigham and Hey 1981) to quantify more precisely the timing, frequency,
duration, and intensity of exposures which cause shifts in community
structure. The differences in toxicity and exposure patterns between the
bluegill and carp reaches on the DuPage River were too great to determine, for
example, whether the community structure characteristic of the bluegill reach
would degrade if the index exceeded 0.3 for brief periods.
This type of information is useful in pollution control engineering,
where a performance standard is achieved within some specified variation and
failure rate. If the biological consequences of excursions beyond the mean
can be specified, then waste control programs can be designed to stay within
the limits without incurring unnecessary expense to achieve lower variation or
failure rates.
The timing of excursions also should be examined, e.g. do excursions in
the spring when larval fish are present have greater effect on fish
populations than the same excursions in late summer? If so, the waste loading
could be adjusted seasonally to protect aquatic life in the stream. Although
the present index, which is based on toxicity to adult fish, can be related
empirically to the status of fish populations in streams, as described in the
above applications, another approach would be to develop toxicity indices for
sensitive life history stages and use them at the appropriate season.
The present index is based on concentrations of toxicants in water. In
the Illinois River, water quality has improved without a concomitant recovery
of infaunal macroinvertebrate communities, because of an apparent legacy of
toxicants remaining in the sediments (Sparks 1984). Toxicity indices should
be developed for reference species representing several trophic levels and
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functional groups, used together to increase the reliability of simulations
and estimations, and verified by field trials or by application to data-rich
environments, as described in this paper.
In particular, indices should be developed for assessing the quality of
sediments, as well as quality of water, using benthic macroinvertebrates and
rooted aquatic macrophytes as reference organisms. Sediment LC50s (in mg of
toxicant per kg of "standard" sediment materials, e.g. montmorillonite clay or
natural sediments of consistent composition) could be determined by adding
reagent grade toxicants to a sediment slurry, allowing it to settle, then
adding the test organisms. The additive assumption should be tested with
mixtures of toxicants in sediments. The database on sediment LC50s would be
employed in a sediment toxicity index just as described above for the water-
based index.
SUMMARY
The toxicity index provides a way of relating water quality monitoring
data to toxicity data available in the literature and to fish populations in
streams, so that stream reaches can be classified according to their
suitability for fish communities of varying sensitivity to common pollutants
and environmental stressors. The index goes beyond classification however, to
identification of causative factors. Chemical concentration units are
converted to toxicity units, so the chemicals which contribute the most
toxicity in a stream reach can be identified. The component toxicities also
can be summed to provide an estimate of the total toxicity in a reach. The
assumption of additive effects appears generally valid if lethal thresholds,
rather than survival times, are used to measure toxicity. Predicted lethal
thresholds are generally within 50% of measured thresholds in laboratory and
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field experiments where complex mixtures are present.
Where sufficient data exist, the effects of factors which modify chemical
equilibria or the sensitivity of fish (temperature, pH, dissolved oxygen,
calcium concentration or hardness) can be taken into account in the
algorithms. One indirect, beneficial result of this systematic search for
information on interactions is that toxicological data gaps and research needs
are identified and prioritized.
Application of the toxicity index to 1970s water quality data from the
Illinois River indicates that ammonia from the Chicago area is a major
contributor to toxicity in the upper river, with sporadic contributions from
cyanide downstream of Peoria. Mean index values above 0.2 occur in upstream
reaches where there are depauperate fish communities dominated by introduced
carp and goldfish. Native fish increase downstream where mean and "worst
case" index values are generally below 0.2. The index also was used in an
after-the-fact analysis to determine that a 1988 spill of urea ammonium
nitrate into the upper Illinois River probably killed fish several miles
downstream.
Toxicity indices in the range 0.2-0.4 have been established as the
threshold for alteration of the species composition of fish communities by
results of field tests in the Illinois River and in several English rivers and
by analysis of toxicity simulations in 3 reaches of the Dupage River: a
fishless reach, a carp-dominated reach, and a reach with a mixed community
including the native bluegill sunfish. A time dimension was added to this
threshold in the DuPage River study, where the index never exceeded a
sublethal value of 0.30 for even 1 hour in the bluegill reach, there were only
2 excursions between 0.25 and 0.30 (lasting an average of 36.5 hours), and the
average toxicity was 0.115 during a simulated 3-year period.
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Toxicity indices are not designed to assess bioconcentration effects or
to be used in place of direct in-plant or in-stream toxicity testing programs.
The projects described here, however, have demonstrated that these indices can
be used to complement water quality monitoring programs by providing numerical
values that describe a biological parameter (toxicity). The recent regulatory
emphasis that has been placed on effluent toxicity testing and biological
monitoring suggests that the results of water quality monitoring programs are
of limited value in assessing toxicity problems. Although it is true that a
limited number of in-stream measurements for a particular toxicant should not
be used in sole support of any important water management decision, the
alternative of not considering these data at all seems equally unacceptable
and, in fact, undermines a common objective of water quality monitoring
programs, which is to provide information on which to base management
decisions.
Unfortunately, the products of most water quality monitoring programs are
voluminous tables of data, which decision makers find difficult to interpret.
Analysis usually is confined to the number of times standards for individual
constituents were exceeded rather than to interactions and their biological
consequences. Toxicity indices provide a logical way to assess the joint
action of toxicants and the modifying effects of environmental factors on
aquatic organisms.
As demonstrated by the projects described here, toxicity indices can be
used to determine which chemical components contribute the most toxicity at a
given location or time, to relate temporal and spatial variations in water
quality to fish community structure, to evaluate alternative pollution control
strategies, to assess the biological effects of spills, and to classify stream
reaches according to their suitability for fish.
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ACKNOWLEDGMENTS
The Illinois River application was supported by the Office of Water
Research and Technology, Project No. A-067-ILL, and by the Illinois Institute
for Natural Resources, Project No. 20.107. Kevin Anderson and Yip Tai-Sang
wrote the computer programs for this project.
The DuPage River application was supported by the U.S. Environmental
Protection Agency, Grant No. R805614010, and was conducted jointly by Dr.
Warren Brigham, Illinois Natural History Survey, and Donald Hey, consultant to
the Northeastern Illinois Planning Commission and Hydrocomp, Inc.
Dr. Kenneth S. Lubinski, currently with the U. S. Fish and Wildlife
Service in LaCrosse, Wisconsin, contributed substantially to various projects
involving toxicity indices from 1973 to 1974, and again from 1979 to 1987.
K. Douglas Blodgett, Illinois Natural History Survey, River Research
Laboratory, Havana, wrote a program in LOTUS to calculate ammonia component
toxicity, using updated information, for analysis of the 3 January 1988
ammonia spill in the Illinois River.
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LITERATURE CITED
Anderson, K.B., R.E. Sparks, and A. A. Paparo. 1978. Rapid assessment of
water quality using the fingernail clam, Musculium transversum.
Illinois Water Resources Center Report No. 133, University of
Illinois. 1 15 p.
Brigham, W., and D. Hey. 1981. A stress function for evaluating
strategies for water quality management. Contract Report. U.S.
Environmental Protection Agency Grant No. R805614010. 92 pp.
Brockway, D. 1950. Metabolic products and their effects. Progressive
Fish Culturist 12:127-129.
Brown, V.M., D.H.M. Jordan, and B.A. Tiller. 1969. The acute toxicity
to rainbow trout of fluctuating concentrations and mixtures of
ammonia phenol and zinc. Journal of Fisheries Biology 1:1-9.
Brown, V. M, D. G. Shurben, and D. Shaw. 1970. Studies on water
quality and the absence of fish from some polluted English rivers.
Water Research 4:363-382.
Cairns, J., Jr. 1957. Environment and time in fish toxicity.
Industrial Wastes 1:1-15.
Carlander, K. D. 1977. Handbook of freshwater fishery biology. Vol. 2.
The Iowa State University Press, Ames, Iowa. 431 p.
Doudoroff, P. 1976. Toxicity to fish of cyanides and related compounds,
a review. U. S. Environmental Protection Agency Ecological Research
Series No. EPA-600/3-76-038. 155 p.
Doudoroff, P., G. Leduc, and C.R. Schneider. 1966. Acute toxicity to
fish of solutions containing complex metal cyanides, in relation to
concentrations of molecular hydrocyanic acid. Transactions of the
American Fisheries Society 95:6-22.
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Toxicity index July 14, 1988 Page 26
Dreher, D.W. 1981. Study of fish toxicity in the East Branch DuPage
River. Report. Northeastern Illinois Planning Commission.
Edwards, R.W., and V.M. Brown. 1966. Pollution and fisheries: a
progress report. Water Pollution Control 66:63-78.
Emerson, K., R.C. Russo, R.E. Lund, and R>V> Thurston. 1975. Aqueous
ammonia equilibrium calculations: effect of pH and temperature.
Journal of the Fisheries Research Board of Canada 32:2379-2383.
Esvelt, L. A., W. J. Kaufman, and R. E. Selleck. 1971. Toxicity removal
from municipal wastewaters. SERL Report No. 71-1. Sanitary
Engineering Research Laboratory, College of Engineering and School
of Public Health, University of California at Berkeley. 224 p.
Fromm, P.O. 1970. Effect of ammonia on trout and goldfish. Pages 9-22
in Toxic Action of Water Soluble Pollutants on Freshwater Fish.
Report No. 18050 DST 12/70. U. S. Environmental Protection Agency,
Water Quality Office, Washington, D.C.
Havera, S.P., F.C. Bellrose, H.K. Archer, F.L. Paveglio, Jr., D.W.
Steffeck, K.S. Lubinski, R.E. Sparks, W.U. Brigham, L. Coutant, S.
Waite, and D. McCormick. 1980. Projected effects of increased
diversion of Lake Michigan water on the environment of the Illinois
River valley. U.S. Army Corps of Engineers, Chicago District. 550
P-
Herbert, D.W.M. 1962. The toxicity to rainbow trout of spent still
liquors from the distillation of coal. Annals of Applied Biology
50:755-777.
Herbert, D.W.M, D.H.M. Jordan, and R. Lloyd. 1965. A study of some
fishless rivers in the industrial Midlands. Journal of the
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Toxicity index July 14, 1988 Page 27
Proceedings of the Institute of Sewage Purification 6:569-582.
Hey, D.L., E.L. Hardin, D.W. Dreher, and N.S. Philippi. 1982. Proposed
revision to the water quality standards for the DuPage River.
Report. Northeastern Illinois Planning Commission. 88 p.
Hey, D.L., J.M. Pappas, and L.C. Cox. 1980. An economic analysis of
effluent standards for BOD, ammonia, total suspended solids, and
disinfection: case study of a modern treatment plant. Document No.
80/25. Illinois Institute of Natural Resources, Environmental
Management Division, Chicago. 46 p.
Jackson, H.O., and W.C. Starrett. 1959. Turbidity and sedimentation at
Lake Chautauqua, Illinois. Journal of Wildlife Management 23:157-
168.
Lloyd, R. 1961. The toxicity of mixture of zinc and copper sulphates to
rainbow trout (Salmo gairdnerii Richardson). Annals of Applied
Biology 49:535-538.
Lloyd, R. 1965. Factors that affect the tolerance of fish to heavy
metal poisoning. Pages 181-187 in C. M. Tarzwell, ed. Biological
problems in water pollution, third seminar. U.S. Department of
Health, Education, and Welfare, Public Health Service, Division of
Water Supply and Pollution Control, Cincinnati, Ohio.
Lloyd, R., and D.H.M. Jordan. 1964. Predicted and observed toxicities
of several sewage effluents to rainbow trout: a further study.
Journal of the Proceedings of the Institute of Sewage Purification,
Pt. 2, pp. 183-186.
Lubinski, K.S. 1981. Modification of a bluegill toxicity index system
for use by the Illinois Environmental Protection Agency: Phase II.
Bluegill toxicity index systems description and protocol
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Toxicity index July 14, 1988 Page 28
development. Contract Report to the Illinois Environmental
Protection Agency, Springfield. 80 p.
Lubinski, K.S. and R. E. Sparks. 1981. Use of bluegill toxicity indexes
in Illinois. Pp. 324-337 in D.R. Branson and K.L. Dickson, ed.
Aquatic Toxicology and Hazard Assessment: Fourth Conference. ASTM
Special Technical Publication No. 737. American Society for Testing
and Materials, Philadelphia, Pennsylvania. 471 p.
Lubinski, K.S., R.E. Sparks, and L.A. Jahn. 1974. The development of
toxicity indices for assessing the quality of the Illinois River.
Illinois Water Resources Center Research Report No. 96, University
of Illinois, Urbana-Champaign, Illinois. 46 p.
Macaitis, W., J. Variakojis, and R. Kuhl. 1987. Water quality proposal.
Metropolitan Sanitary District of Greater Chicago. 86 p., 7
appendices.
Marvin, D.E., and A.G. Heath. 1968. Cardiac and respiratory responses
to gradual hypoxia in three ecologically distinct species of
freshwater fish. Comparative Biochemistry and Physiology 27:349-
355.
Mills, H.B., W.C. Starrett, and F.C. Bellrose. 1966. Man's effect on
the fish and wildlife of the Illinois River. Illinois Natural
History Survey Biological Notes No. 57. 24 p.
Merkens, J. C. and K. M. Downing. 1957. The effect of tension of
dissolved oxygen on the toxicity of un-ionized ammonia to several
species of fish. Annals of Applied Biology 45:521-527.
Muchmore, C.B., W.M. Lewis, R.C. Heidinger, M.H. Paller, and L.J.
Wawronowicz. 1979. Impact of the existing ammonia nitrogen waste
quality standard. Illinois Institute of Natural Resources Project
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Toxicity index July 14, 1988 Page 29
Nos. 80.137, 80.138, and 80.153.
Neuhold, J. M. and W. F. Sigler. 1960. Effects of sodium fluoride on
carp and rainbow trout. Transactions of the American Fisheries
Society 89:358-370.
Reinbold, K.A., and S.M. Pescitelli. 1981. Effects of cold temperature
on toxicity of ammonia to rainbow trout, bluegills and fathead
minnows. Contract Report. Contract No. 68-01-5832. U. S.
Environmental Protection Agency, Region V, Water Division, Chicago,
Illinois. 25 p.
Reinbold, K.A., and S.M. Pescitelli. 1982. Effects of exposure to
ammonia on sensitive life stages of aquatic organisms. Contract
Report. Contract No. 68-01-5832. U. S. Environmental Protection
Agency, Region V, Water Division, Chicago, Illinois. 33 p.
Roseboom, D.P. and D.L. Richey. 1977. Acute toxicity of residual
chlorine and ammonia to some native Illinois fishes. Report of
Investigation 85. Illinois State Water Survey, Urbana, Illinois.
42 p.
Sparks, R.E. 1984. The role of contaminants in the decline of the
Illinois River: implications for the Upper Mississippi. Pages 25-65
in J.G. Wiener, R.V. Anderson, D.R. McConville, eds. Contaminants in
the Upper Mississippi River. Butterworth Publishers, Stoneham,
Massachusetts. 384 p.
Sparks, R.E. and K.B. Anderson. 1977. Toxicity of ammonia in mixtures
and development of a toxicity index for use in a stream
classification system for Illinois: summary report. Pages 34-40 in.
B.B. Ewing, ed. Feasibility of a systematic approach to water
quality management in Illinois. Report of the Stream/Lake
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Toxicity index July 14, 1988 Page 30
Classification Project. Illinois Institute for Environmental
Quality Document No. 77/35.
Sparks, R.E., and W.C. Starrett. 1975. An electrofishing survey of the
Illinois River, 1959-1974. Illinois Natural History Survey Bulletin
31(8):317-380.
Sprague, J. B. and B. A. Ramsay. 1965. Lethal levels of mixed copper-
zinc solutions for juvenile salmon. Journal Fisheries Research
Board of Canada 22:425-432.
Sprague, J.B. 1970. Measurement of pollutant toxicity to fish. 1.
Bioassay methods for acute toxicity. Water Research 4:3-32.
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FIGURE LEGENDS AND LIST OF TABLES FOR TOXICITY INDEX MANUSCRIPT BY S
Figure 1. Mean and maximum toxicity indices in the Illinois Waterway, 1972-
1974. River mileages begin downstream at the confluence with the
Mississippi, at river mile 0, and progress upstream toward Chicago
and Lake Michigan, at river mile 330. The horizontal line is drawn
at a toxicity index value of 0.2, below which populations of native
species can maintain themselves, if factors other than acute
toxicity are not limiting. An index of 1.0 is lethal, equivalent
to the 96-hour LC50. See text for explanation.
Figure 2. Cyanide and ammonia component toxicities in the Illinois Waterway,
1972-1974. River mileages begin downstream at the confluence with
the Mississippi, at river mile 0, and progress upstream toward
Chicago and Lake Michigan, at river mile 330. The horizontal line
is drawn at a toxicity index value of 0.2, below which populations
of native species can maintain themselves, if factors other than
acute toxicity are not limiting. An index of 1.0 is lethal,
equivalent to the 96-hour LC50. See text for explanation.
Figure 3. Toxicity indices at 1-hour intervals in 3 reaches of the DuPage
River on 17 March 1972. Indices are output from a simulation
model, calibrated for the DuPage River.
Table 1. Toxicity of a total ammonia-N concentration of 47 mg/1 in the
Illinois River under different conditions.
Table 2. The number of excursions past different toxicity levels in the
DuPage River during a simulated 3-year period.
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Table 3. Impact of a water quality management plan on toxicity in the DuPage
River.
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TABLE 1 —Toxicity of a total ammonia-N concentration of 47 mg/1 to 1-gram
bluegills under different conditions of temperature (°C), pH, and
dissolved oxygen (DO).
PH
DO
Bluegill Toxicity
Temp
mg/1
% Saturation
Index
28
8.0
7.5
100
2.69
28
8.0
5.0
66
4.83
28
8.0
3.0
40
13.4
28
8.0
1.0
13
3080
28
7.0
5.0
66
0.514
28
7.5
5.0
66
1.60
28
8.0
5.0
66
4.83
28
8.5
5.0
66
13.4
4
8.0
12.2
100
1.90
8
8.0
11.1
100
2.05
12
8.0
10.1
100
2.22
4
8.0
8.1
66
3.43
4
8.0
4.9
40
9.33
4
8.0
1.6
13
534
TABLE 2 — The number of excursions past different toxicity levels
for various lengths of time during a 3-year period,
1 October 1970 to 30 September 1973, in 3 reaches of the
DuPage River (Lubinski 1981).
Num
iber
of To
xicity
Index
Excursions
for 3
Duration Times
Toxicity Index
BGTUs
1 h
24 h
96 h
Fishless reach
1.0
73
51
42
2.0
155
55
34
3.0
194
39
21
Carp reach
1.0
74
31
14
2.0
16
1
3.0
12
1
Bluegill reach
0.1
194
29
15
0.2
37
25
11
0.3
TABLE 3 — Impact of a hypothetical water quality management
plan (see text for explanation) upon the toxicity
index in the fishless reach, DuPage River, DuPage
County, Illinois (Brigham and Hey 1981).
Without
With
Plan
Plan
Toxicity Index
max.
784
51.5
min.
0.120
0.118
mean.
23.0
2.12
ComDonent Toxicity
Ammonia
20.2
0.023
Cyanide
0.088
0.079
Lead
0.023
0.000
Zinc
0.004
0.003
Copper
0.020
0.008
LAS
0.772
0.073
Chlorine
J. 94
1.94
Illinois Waterway
Mean Toxicity Indices
River Mile
Illinois Waterway
Maximum Toxicity Indices
1972
340
River Mile
Illinois Waterway
Cyanide Component Toxicity
340
0.30 -
C 0.20
Illinois Waterway
Ammonia Component Toxicity
River Mile
CO
13
CM
E
N-
'(f)
o>
5—
SI
o
>
CO
be
^
h-
D)
CC5
a.
13
Q
CO
Q
O
o
X
s^iun Ajiojxoi ||iBen|g
Literature Cited
Cairns, J., Jr., R.E. Sparks, and W.T. Waller. 1973. A tentative proposal
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Dickson, K.L., D. Gruber, C. King, and K. Lubenski. 1980. Biological
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