United States Office of Water EPA-822-R-99-014
Environmental Protection 4304 December 1999
Agency
&EPA 1999 Update
of Ambient
Water Quality
Criteria
for
Ammonia
Supersedes 1998 Update
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1999 Update of
Ambient Water Quality Criteria for Ammonia
September 1999
Supersedes 1998 Update
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Washington, D.C.
Office of Research and Development
Mid-Continent Ecology Division
Duluth, Minnesota
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Notices
This update provides guidance to States and Tribes authorized to establish water
quality standards under the Clean Water Act (CWA), to protect aquatic life from acute
and chronic effects of ammonia. Under the CWA, States and Tribes are to establish
water quality criteria to protect designated uses. State and tribal decision makers
retain the discretion to adopt approaches on a case-by-case basis that differ from this
guidance when appropriate. While this update constitutes EPA's scientific
recommendations regarding ambient concentrations of ammonia that protect freshwater
aquatic life, this update does not substitute for the CWA or EPA's regulations; nor is it a
regulation itself. Thus, it cannot impose legally binding requirements on EPA, States,
Tribes, or the regulated community, and might not apply to a particular situation based
upon the circumstances. EPA may change this guidance in the future.
This document has been approved for publication by the Office of Science and
Technology, Office of Water, U.S. Environmental Protection Agency. Mention of trade
names or commercial products does not constitute endorsement or recommendation for
use.
Acknowledgment
This update is a modification of the 7998 Update of Ambient Water Quality Criteria for
Ammonia. The 1999 modifications were written by Charles Delos and Russ Erickson.
The 1998 Update, most of the text of which was carried unchanged into the 1999
Update, was written by Charles Stephan, Russ Erickson, Charles Delos, Tom
Willingham, Kent Ballentine, and Rob Pepin (with substantial input from Alex Barren of
the Virginia Department of Environmental Quality, Dave Maschwitz of the Minnesota
Pollution Control Agency, and Bob Mosher of the Illinois Environmental Protection
Agency) under the auspices of the Aquatic Life Criteria Guideline Committee. Please
submit comments or questions to: Charles Delos, U.S. EPA, Mail Code 4304,
Washington, DC 20460 (e-mail: delos.charles@epa.gov).
IV
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Contents
Page
Notices and Acknowledgment ii
Introduction 1
Overview of Ammonia Toxicology 2
Temperature-Dependence of Ammonia Toxicity 8
pH-Dependence of Ammonia Toxicity 28
Formulation of the CMC 36
Review and Analysis of Chronic Data 43
Seasonality of Chronic Toxicity Endpoints 68
Formulation of the CCC 74
Chronic Averaging Period 81
The National Criterion for Ammonia in Fresh Water 83
References 89
Appendix 1: Review of Some Toxicity Tests 101
Appendix 2: Methods for Regression Analysis of pH Data 104
Appendix 3: Conversion of Results of Toxicity Tests 107
Appendix 4: Acute Values 109
Appendix 5: Histopathological Effects 118
Appendix 6: Results of Regression Analyses of Chronic Data 122
Appendix 7: Acute-Chronic Ratios 135
Appendix 8: A Field Study Relevant to the CCC 139
Appendix 9: Water-Effect Ratios 147
V
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VI
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Introduction
Since the U.S. EPA published "Ambient Water Quality Criteria for Ammonia -1984"
(U.S. EPA 1985a), it has issued additional information on aquatic life criteria for
ammonia (Heberand Ballentine 1992; U.S. EPA 1989,1996, 1998). Also, results of
additional toxicity tests on ammonia have been published since 1985, which could
affect the freshwater criterion for ammonia. The purpose of the 1998 Update was to
revise the 1984/1985 ammonia criteria document (U.S. EPA 1985a) and replace Heber
and Ballentine (1992) and U.S. EPA (1996) by addressing selected important issues to
the extent possible in a short-term effort without additional research.
EPA obtained public comment on the 1998 Update. In response to those comments
EPA modified its criteria recommendations and prepared the current document, the
1999 Update, which replaces and supersedes the 1998 Update. A supporting
document, Response to Comments on 1998 Update of Ambient Water Quality Criteria
for Ammonia (U.S. EPA 1999) presents the comments and EPA's responses.
This 1999 Update first presents an overview of ammonia toxicology in order to provide
the background needed to explain the revisions of the freshwater ammonia criterion.
Then the equations used in the older documents to address the temperature- and pH-
dependence of ammonia toxicity in fresh water are revised to take into account newer
data, better models, and improved statistical methods. Next, a CMC (acute criterion) is
derived from the acute toxicity data in the 1984/1985 criteria document, pH-normalized
using the new equations. Then, new and old chronic toxicity data are evaluated and
used to derive a CCC (chronic criterion). Lastly, the chronic averaging period is
addressed.
The 1999 Update differs from the 1998 Update primarily in the handling of the
temperature-dependency for the CCC, and therefore the formulation of the CCC, and
the expression of the national criterion. One facet of the chronic averaging period was
also slightly changed. Neither the 1998 Update nor the 1999 Update were intended to
address (1) the CMC averaging period, or (2) the frequency of allowed exceedances.
The Updates address only the freshwater criterion and do not affect the saltwater
criterion for ammonia (U.S. EPA 1989).
Concentrations of un-ionized ammonia and total ammonia are given herein in terms of
nitrogen, i.e., as mg N/L, because most permit limits for ammonia are expressed in
terms of nitrogen. CMCs and CCCs are given to three significant figures to minimize
the effect of round-off error in the calculation of permit limits.
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Overview of Ammonia Toxicology
The 1984/1985 ammonia criteria document reviewed data regarding the dependence of
the toxicity of ammonia to aquatic organisms on various physicochemical properties of
the test water, especially temperature, pH, and ionic composition. A key factor in these
relationships is the chemical speciation of ammonia. In aqueous solution, ammonia
primarily exists in two forms, un-ionized ammonia (NH3) and ammonium ion (NH4+),
which are in equilibrium with each other according to the following expressions:
NH4 * NH3 + H+ (1)
„
K =
[NH4]
(2)
The equilibrium constant K depends significantly on temperature; this relationship has
been described by Emerson et al. (1975) with the following equation:
pK = 0.09018 + • (3)
273.2 + T { '
where pK = -log10K and T is temperature in degrees Celsius.
From Equation 2, the definition of pK, and the definition
pH = -log10[H+], the following expressions can be derived for the fraction of total
ammonia in each of the two forms:
NH3
(4)
The individual fractions vary markedly with temperature and pH. The pH-dependence
of the relative amounts of un-ionized ammonia and ammonium ion at 25°C, at which
pK=9.24, is illustrated in the following graph:
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Chemical Speciation of Ammonia
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Ammonia speciation also depends on ionic strength, but in fresh water this effect is
much smaller than the effects of pH and temperature (Soderberg and Meade 1991) and
is sufficiently small compared to the typical uncertainty in LCSOs that it will not be
considered here as a variable affecting ammonia toxicity. (As discussed later, ionic
composition might affect ammonia toxicity in ways other than its effect on ammonia
speciation).
These speciation relationships are important to ammonia toxicity because un-ionized
ammonia is much more toxic than ammonium ion. The importance of un-ionized
ammonia was first recognized when it was observed that increased pH caused total
ammonia to appear to be much more toxic (Chipman 1934; Wuhrmann and Woker
1948). It is not surprising that un-ionized ammonia is the more toxic form, because it is
a neutral molecule and thus is able to diffuse across the epithelial membranes of
aquatic organisms much more readily than the charged ammonium ion. Ammonia is
unique among regulated pollutants because it is an endogenously produced toxicant
that organisms have developed various strategies to excrete, which is in large part by
passive diffusion of un-ionized ammonia from the gills. High external un-ionized
ammonia concentrations reduce or reverse diffusive gradients and cause the buildup of
ammonia in gill tissue and blood.
Because of the importance of un-ionized ammonia, it became a convention in the
scientific literature to express ammonia toxicity in terms of un-ionized ammonia, and
water quality criteria and standards followed this convention. However, there are
reasons to believe that ammonium ion can contribute significantly to ammonia toxicity
under some conditions. Observations that ammonia toxicity is relatively constant when
expressed in terms of un-ionized ammonia come mainly from toxicity tests conducted at
pH>7.5. At lower pH, toxicity varies considerably when expressed in terms of un-
ionized ammonia and under some conditions is relatively constant in terms of
ammonium ion (Erickson 1985). Also, studies have established that mechanisms exist
for the transport of ammonium ion across gill epithelia (Wood 1993), so this ion might
contribute significantly to ammonia exchange at gills and affect the buildup of ammonia
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in tissues if its external concentration is sufficiently high. Thus, the very same
arguments employed for the importance of un-ionized ammonia can also be applied in
some degree to ammonium ion. This is not to say that ammonium ion is as toxic as un-
ionized ammonia, but rather that, regardless of its lower toxicity, it can still be important
because it is generally present in much greater concentrations than un-ionized
ammonia.
Also, when expressed in terms of un-ionized ammonia, ammonia toxicity is usually not
constant with temperature, on average being about four-fold greater at 5°C than at
25°C for fish (Erickson 1985). Because the relative amount of ammonium ion is also
higher at low temperatures, this raises the possibility that ammonium ion might be in
part responsible for this temperature dependence. However, temperature might also
alter ammonia toxicity by affecting membrane permeability, endogenous ammonia
production, and other physiological processes.
Various authors have evaluated models that might explain the pH and temperature
dependence of ammonia toxicity. Tabata (1962) and Armstrong et al. (1978) suggested
that the observed pH dependence is due to joint toxicity of un-ionized ammonia and
ammonium ion.
Toxicity of Ammonia
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Ammonia
nia .... /SvTTTTT-:
The adjacent graph shows an
idealized picture of ammonia
toxicity assuming that (a)
ammonium ion and un-
ionized ammonia jointly
determine toxicity and (b) un-
ionized ammonia is 100 times
more toxic than ammonium
ion. At sufficiently high pH,
the more toxic un-ionized
ammonia comprises a
sufficiently large fraction of
total ammonia to dominate
toxicity, and so toxicity is
relatively constant when
expressed in terms of un-
ionized ammonia. As pH
decreases, the relative amount of ammonium ion increases until it contributes
significantly to toxicity, so that toxicity expressed in terms of un-ionized ammonia
increases (i.e., it appears that less un-ionized ammonia is necessary to cause toxicity
because ammonium ion is responsible for some of the toxicity). At sufficiently low pH,
ammonium ion dominates toxicity, and so toxicity is relatively constant when expressed
in terms of either ammonium ion or total ammonia.
PH
In contrast to this theory, Lloyd and Herbert (1960) suggested that the apparent effect
of pH on un-ionized ammonia toxicity is due to the data being plotted in terms of the pH
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of the bulk exposure water rather than the pH at the gill surface. The release of carbon
dioxide at the gill lowers pH when pH is moderately alkaline, but has less effect when
pH is already low; this results in an apparent effect of pH on toxicity when the pH of the
bulk exposure water is used even if there is no such effect if the pH at the gill surface is
used. Szumski et al. (1982) suggested that this theory explained not only much of the
pH dependence of ammonia toxicity, but also the temperature dependence.
Erickson (1985) reviewed available information concerning the effects of pH and
temperature on acute toxicity of ammonia when expressed in terms of un-ionized
ammonia and tested its adherence to these theories. He concluded that effects
associated with pH changes at the gill could not account for the effect of temperature
and only a small part of the effect of pH. In contrast, the additive joint toxicity model
explained a large part of the dependence of ammonia toxicity on pH and predicted
important features of the data, specifically a slope of zero at high pH and a slope of one
at low pH. The joint toxicity model could also be fit to the temperature data, but led to
values of the model parameters that were questionable because they indicated that
ammonium ion is as toxic or more toxic than un-ionized ammonia. Clearly, joint toxicity
could not possibly account for both pH and temperature effects, and Erickson (1985)
concluded that joint toxicity is likely responsible for much of the pH effect, but not for
the temperature effect. In the 1984/1985 criteria document, it was noted that the one
available dataset concerning the dependence of chronic toxicity on pH (Broderius et al.
1985) also suggested joint toxicity of un-ionized ammonia and ammonium ion.
Therefore, a major consideration in deriving the aquatic life criterion for ammonia is
whether the mathematical model used to describe pH dependence should be based on
joint toxicity theory. Since the 1984/1985 criteria document was issued, several
additional studies (Sheehan and Lewis 1986; Schubauer-Berigan et al. 1995; Ankley et
al. 1995; Johnson 1995) of the pH dependence of ammonia toxicity have provided more
information regarding the relative importance of un-ionized ammonia and ammonium
ion, including indications of more diversity of pH behavior among species than was
apparent in the data reviewed by Erickson (1985).
The report of Sheehan and Lewis (1986) requires special consideration here because
they suggest that the toxicity of ammonia at low pH is due to the effect of osmotic shock
on unacclimated organisms and that this has major implications for the derivation of a
criterion for ammonia. In their investigations concerning the pH-dependence of acute
ammonia toxicity to channel catfish, Sheehan and Lewis (1986) found that LCSOs
expressed in terms of un-ionized ammonia increased with increasing pH, but less so
than reported in most studies, although Tomasso et al. (1980) also reported little effect
of pH>7 on un-ionized ammonia toxicity to the channel catfish. Sheehan and Lewis
noted that lethal concentrations at pH=6 were associated with very high total ammonia
concentrations (2000 mg N/L) and exhibited steeper concentration-effect curves than at
higher pH. They also reported that other salts were lethal at similar concentrations and
suggested that the toxicity of ammonia at low pH was due to the effect of osmotic shock
on unacclimated organisms rather than a specific action of the ammonium ion per se.
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However, the implication of this work for the ammonia criterion is doubtful for the
following reasons:
1. Any concern that the effects of high concentrations of ammonia would be less for
acclimated organisms is really not relevant. To be adequately protective, criteria
cannot assume that acclimation takes place, because if such high ammonia
concentrations are discharged, they would create a plume of high concentrations
compared to ambient levels. Organisms entering that plume would not be
acclimated to the high concentrations.
2. It is doubtful that the effects of high salt concentrations observed by Sheehan and
Lewis were strictly due to osmotic effects. In their experiments, potassium chloride
caused higher mortality than the physiologically balanced salt they also used. In
fact, the toxicities of such salts vary quite widely, with potassium salts generally
being more toxic (Mount et al. 1997), probably due to effects of potassium beyond
any osmotic effects. Ammonium chloride also caused higher mortality than the
physiologically balanced salt, although this might be in part due to effects of un-
ionized ammonia.
3. As part of their evidence for supporting osmotic effects as a toxic mechanism at low
pH, Sheehan and Lewis noted that the dose-response curves were steeper at low
pH, suggestive not only of a different mechanism, but one that is less variable
among organisms within a test. However, Broderius et al. (1985) found the
opposite effect of pH on dose-response curves.
4. The LCSOs for channel catfish at low pH are generally much higher than those for
other fishes that have been tested at low pH. When expressed in terms of total
ammonia, the LC50 for channel catfish at pH=6 is four-fold higher than any other
LC50 reported for a fish species. For many other fishes, LCSOs at pH«6.5
represent salt concentrations of only a few hundred mg/L and less than a factor of
two greater than that of control water. A role of osmotic effects in such cases is
doubtful. Of all of the fish species tested, the pH curves for channel catfish show
the least indication for an effect of ammonium ion, so it is a very questionable
species upon which to base broad conclusions.
5. In contrast to Sheehan and Lewis, Knoph (1992) reported no mortality of Atlantic
salmon at pH=6 in KCI or in physiologically balanced salt solutions with
concentrations equivalent to ammonium chloride solutions causing 45% mortality.
Similarly, Mount et al. (1997) found acute LCSOs for fathead minnows for various
salts and combinations (except those including potassium) to be at least several-
fold higher than the total ammonia LCSOs reported at pH=6.5 by Thurston et al.
(1981 b). Although for an invertebrate, the likely role of ammonium ion other than in
association with high salt concentrations is also evident in the daphnid data of
Tabata (1962) and Mount et al. (1997).
6. Even if a different mechanism for toxicity exists at low pH, these tests still identify
concentrations that are unacceptably toxic and this is still joint toxicity in the broad
sense of the term. Although the joint toxicity might not be strictly additive, as would
be expected if the two forms of ammonia operate by the same mechanism, it is joint
toxicity nonetheless and should exhibit a similar pH dependence and be
considered in criteria derivation.
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Although there is considerable reason to consider the effects of pH on ammonia toxicity
to be largely due to the joint toxicity of ammonium ion and un-ionized ammonia, pH can
have other effects on membrane function and other physiological processes that could
also alter ammonia toxicity, especially at very low and high pHs, and these are poorly
established. The state of knowledge for the pH dependence is incomplete in terms of
understanding specific mechanisms, variation among species, and interactions with
various physicochemical processes. Lacking a definitive, thorough theoretical
approach for describing pH effects, the most reasonable approach is to adopt the best
empirical description that can be obtained from available data. However, the shape of
this empirical equation can be guided by consideration of the evidence for the role of
speciation in ammonia toxicity.
The effects of temperature on ammonia toxicity are even less well understood, and
there is no adequate theoretical basis or scientific understanding for specifying how
temperature adjustments to the ammonia criterion should be made. Therefore, an
empirical approach will also be used for temperature dependence, as developed in the
next section.
As reviewed in the 1984/1985 ammonia criteria document, ammonia toxicity can also
depend on various aspects of the ionic composition of the exposure water, but the
effects were not clear and consistent enough to warrant inclusion of other variables in
the criterion. Although Soderberg and Meade (1992), Yesaki and Iwama (1992),
Ankley et al. (1995), Johnson (1995), Borgmann and Borgmann (1997), and Iwama et
al. (1997) have provided new data concerning interactions between various ions and
ammonia toxicity and excretion, there is still insufficient understanding and information
to account for these effects in the criterion and they will have to be addressed using
water-effect ratios or other site-specific approaches.
In summary, the available evidence indicates that the toxicity of ammonia can depend
on ionic composition, pH, and temperature. The mechanisms of these effects are
poorly understood, but the pH dependence strongly suggests that joint toxicity of un-
ionized ammonia and ammonium ion is an important component. For the reasons
presented above, the following approach will be used to account for these effects.
1. Because the effects of ionic composition on ammonia speciation in fresh water are
small and its other effects on toxicity are poorly established, the ionic composition
of the exposure water will not be considered in the derivation of the criterion.
2. Even though temperature can strongly affect the relative amounts of un-ionized
ammonia and ammonium ion, its effect on the toxicity of ammonia is not strongly
indicative of joint toxicity and will be described strictly by an empirical approach.
3. The effect of pH will be described by equations that include basic features of joint
toxicity of un-ionized ammonia and ammonium ion, but with an empirical
component that recognizes the incomplete knowledge of these effects.
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Temperature-Dependence of Ammonia Toxicity
This section presents the temperature analysis published in the 1998 Update, followed
by the re-analysis performed for this 1999 Update.
1998 Analysis of Temperature-Dependence
The 1984/1985 ammonia criteria document identified temperature as an important
factor affecting the toxicity of ammonia. When expressed in terms of un-ionized
ammonia, the acute toxicity of ammonia was reported in the criteria document to be
inversely related to temperature for several species offish, whereas limited data on
acute ammonia toxicity to invertebrates showed no significant temperature
dependence. No direct data were available concerning the temperature dependence of
chronic toxicity. It was noted, however, that the differences between chronic values for
salmonid fish species tested at low temperatures and chronic values for warmwater fish
species tested at higher temperatures paralleled differences in acute toxicity known to
be caused by temperature.
In the 1984/1985 criteria document, an average temperature relationship observed for
fish was used to adjust fish acute toxicity data to a common temperature (20°C) for
derivation of the CMC for un-ionized ammonia; this same relationship was used to
extrapolate this CMC to other temperatures. (Invertebrate toxicity data were not
adjusted, but invertebrates were sufficiently resistant to ammonia that adjustment of
invertebrate data was not important in the derivation of the CMC.) This temperature
relationship for fish resulted in the un-ionized ammonia CMC being higher at warm
temperatures than at cold temperatures. Additionally, because of concerns about the
validity of extrapolating the temperature relationship to high temperatures, the un-
ionized ammonia CMC was "capped" to be no higher than its value at a temperature,
called TCAP, near the upper end of the temperature range of the acute toxicity data
available for warmwater and coldwater fishes. Similarly, the CCC was capped at a
temperature near the upper end of the temperature range of the available chronic
toxicity data.
Although the un-ionized ammonia criterion is lower at low temperatures, this does not
result in more restrictive permit limits for ammonia because the ratio of ammonium ion
to un-ionized ammonia increases at low temperatures, resulting in the total ammonia
criterion being essentially constant at temperatures below TCAP. In practice, however,
the criterion at low temperatures can be more limiting for dischargers than the criterion
at high temperatures because biological treatment of ammonia is more difficult at low
temperatures. Above TCAP, the constant un-ionized ammonia criterion results in the
total ammonia criterion becoming progressively lower with increasing temperature,
which can also result in restrictive discharge limitations.
8
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Because more data are available at moderate temperatures than at lower and higher
temperatures, the ammonia criterion is most uncertain for circumstances when
compliance can be most difficult, either because of the low total ammonia criterion at
high temperatures or because of treatment difficulties at low temperatures. This
section examines the data used in the 1984/1985 criteria document and newer data to
determine (1) whether the use of TCAPs should be continued and (2) whether a lower
un-ionized criterion at low temperature is warranted. Data used include those analyzed
by Erickson (1985), which are shown in Figure 2 of the 1984/1985 document, and more
recent data reported by Arthur et al. (1987), DeGraeve et al. (1987), Nimmo et al.
(1989), and Knoph (1992).
Data not used include those reported by the following:
1. Bianchini et al. (1996) conducted acute tests at 12 and 25°C, but one test was in
fresh water, whereas the other was in salt water.
2. Diamond et al. (1993) conducted acute and chronic toxicity tests on ammonia at 12
and 20°C using several vertebrate and invertebrate species. When expressed in
terms of un-ionized ammonia, they reported that vertebrates (i.e., fishes and
amphibians) were more sensitive to ammonia at 12°C than at 20°C, whereas
invertebrates were either less sensitive or no more sensitive at 12°C, compatible
with the relationships used in the 1984/1985 criteria document. However, such
factors as dilution water and test duration varied between tests at different
temperatures and possibly confounded the results (see Appendix 1), raising doubts
about the temperature comparisons for the vertebrates and invertebrates.
Arthur et al. (1987) measured the acute toxicity of ammonia to several fish and
invertebrate species at ambient temperature during different seasons of the year. For
three of the five fish species (rainbow trout, channel catfish, and white sucker), the
relationship of toxicity to temperature was similar to that used in the 1984/1985 criteria
document. When expressed in terms of un-ionized ammonia, no clear relationship
existed between temperature and toxicity for the other fish species (fathead minnow
and walleye). This result for the fathead minnow is different from those of three other
studies (Reinbold and Pescitelli 1982a; Thurston et al. 1983; DeGraeve et al. 1987)
reporting a significant effect of temperature on the acute toxicity of un-ionized ammonia
to the fathead minnow. For five invertebrate species, each tested over a temperature
range of at least 10°C, there was no consistent relationship between temperature and
un-ionized ammonia toxicity. An initial report of these results (West 1985) was the
basis for no temperature adjustment being used for invertebrate data in the 1984/1985
criteria document. Further analysis of the Arthur (1987) data is discussed later.
DeGraeve et al. (1987) studied the effect of temperature (from 6 to 30°C) on the toxicity
of ammonia to juvenile fathead minnows and channel catfish using acute (4-day) and
chronic (30-day) ammonia exposures. As shown for both fish species in Figure 1,
log(96-hr un-ionized ammonia LC50) versus temperature was linear within the reported
uncertainty in the LCSOs; the slopes were similar to those reported in the 1984/1985
criteria document. Problems with the channel catfish chronic tests precluded effective
use of those data and the highest tested ammonia concentrations in the fathead
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minnow chronic tests at 15 and 20°C did not cause sufficient mortality to be useful.
However, sufficient mortality did occur in the fathead minnow chronic tests at 6, 10, 25,
and 30°C. Based on regression analysis of survival versus log concentration
(discussed in more detail in the section concerning the CCC below), 30-day LC20s for
un-ionized ammonia were 0.11, 0.18, 0.48, and 0.44 mg N/Lat6, 10, 25, and 30°C,
respectively. This temperature dependence (Figure 1) is similar to that for acute
toxicity and that used in the 1984/1985 criteria document. The actual effect of
temperature on these 30-day LC20s is probably somewhat greater, because test pH
decreased with increasing temperature.
Nimmo et al. (1989) conducted acute toxicity tests on ammonia at 6 and 20°C in a well
water using Johnny darters and in a river water using both Johnny darters and juvenile
fathead minnows. In all three sets of tests, LCSOs expressed in terms of un-ionized
ammonia were significantly higher at the warmer temperature, by factors ranging from
3.5 to 6.2.
Knoph (1992) conducted acute toxicity tests at temperatures ranging from 2 to 17°C
using Atlantic salmon parr, one series of tests at pH«6.0 and the other at pH«6.4. In
both series of tests, LCSOs expressed in terms of un-ionized ammonia increased
substantially with temperature.
Even with these additional data, the shape of the temperature relationship is not
completely resolved, especially for chronic toxicity. Nevertheless, the acute data for
fishes overwhelmingly indicate that ammonia toxicity, expressed in terms of un-ionized
ammonia, decreases with increasing temperature.
Most importantly, the data of DeGraeve et al. (1987) show (Figure 1) that (a) a linear
relationship of log un-ionized ammonia LC50 versus temperature applies within the
reported uncertainty in the LCSOs over the range of 6 to 30°C and (b) temperature
effects on long-term mortality are similar to those on acute mortality. For invertebrates,
acute toxicity data suggest that ammonia toxicity, when expressed in terms of un-
ionized ammonia, does not decrease, and possibly even increases, with increasing
temperature. Quantifying and adjusting data for this relationship is not necessary
because even at warm temperatures invertebrates are generally more resistant to acute
ammonia toxicity than fishes and thus their precise sensitivities are of limited
importance to the criterion. At low temperatures, they are even more resistant relative
to fishes and thus their precise sensitivity is even less important to the criterion.
Based on this information, the two issues raised above were resolved as follows:
1. TCAPs will not be used in the ammonia criterion. This does not mean that the
notion of high temperature exacerbating ammonia toxicity is wrong; rather, it
reflects the fact that such an effect is not evident in the available data, which cover
a wide temperature range.
2. A CMC, if it were expressed as un-ionized ammonia (rather than total ammonia,
used in this document) would continue to be lower at lower temperatures,
consistent with the observed temperature dependence of ammonia toxicity to the
10
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Figure 1. The effect of temperature on ammonia toxicity in terms of un-ionized ammonia
(DeGraeve et al. 1987). Symbols denote LCSOs or LC20s and 95% confidence
limits and lines denote linear regressions of log LC versus temperature.
4 r
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0.8
0.6
0.4
0.2
0.1
Fathead Minnow
96-hr LC50
5 10 15 20 25 30
Temperature (C)
D)
•| 0.8
o
I 0.6
§ 0.4
c
o
0.2
0.1
Channel Catfish
96-hr LC50
5 10 15 20 25 30
Temperature (C)
D)
E
TO
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T3
0
N
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O
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0.8
0.6
0.4
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Fathead Minnow
30-day LC20
5 10 15 20 25 30
Temperature (C)
11
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most sensitive species, i.e., fishes. Although it is possible that the temperature
relationship differs among fish species and that using the same relationship for all fish
species introduces some uncertainty, specifying a relationship for each fish species is
not possible with current data and would also introduce considerable uncertainty.
Therefore, for a criterion expressed in terms of un-ionized ammonia, available data
support the continued use of a generic temperature relationship similar to that in the
1984/1985 ammonia criteria document, but without TCAPs.
This raised a new issue, however, because the criterion expressed in terms of total
ammonia is nearly constant over all tested temperatures, and the small effect of
temperature on the total ammonia criterion in the 1984/1985 criteria document is
largely an artifact of conducting regression analyses in terms of un-ionized ammonia
and is not indicative of any established, significant trend. It was thought that the
expression and implementation of the ammonia criterion might have been simplified if
temperature were dropped as a modifying factor, which might have been possible if
ammonia toxicity is expressed in terms of total ammonia. Furthermore, permit limits
and compliance are usually expressed in terms of total ammonia nitrogen, and so
expressing the criterion in terms of total ammonia nitrogen would simplify its
implementation by eliminating conversions to and from un-ionized ammonia. Because
of such benefits and because there are no compelling scientific or practical reasons for
expressing the criterion in terms of un-ionized ammonia, the freshwater toxicity data
concerning temperature dependence were reanalyzed in terms of total ammonia
nitrogen.
The data analyzed are from the studies included in the 1984/1985 ammonia criteria
document and the studies of DeGraeve et al. (1987), Nimmo et al. (1989), and Knoph
(1992). All analyses were conducted in terms of total ammonia nitrogen, either as
reported by the authors or as converted by us from reported values for un-ionized
ammonia, pH, and temperature using the speciation relationship of Emerson et al.
(1975). The data are presented in Figure 2 and show considerable diversity, with some
datasets showing decreasing toxicity with increasing temperature, some showing
increasing toxicity, and some showing virtually no change. There are even differences
among studies using the same test species. However, in no case is the effect of
temperature particularly large, being no more than a factor of 1.5 over the range of any
dataset, except for the Johnny darter data of Nimmo et al. (1989). In some studies, test
pH was correlated with test temperature. To reduce the confounding effect of pH, the
total ammonia LC50 was adjusted to the mean pH of the data for the study using the pH
relationship discussed in the next section of the Update. These adjusted data are
shown in Figure 3 and also show neither large effects nor any clear consistency among
or within species or studies.
For each dataset containing at least three data points, a linear regression of log LC50
versus temperature was conducted (Draper and Smith 1981) and the resulting
regression lines are plotted as solid lines in Figures 2 and 3. These regressions are
significant at the 0.05 level for only one dataset (the unadjusted fathead minnow data of
Thurston et al. 1983); for this dataset, however, the regression is not significant when
12
-------
Figure 2. The effect of temperature on acute ammonia toxicity in terms of total ammonia.
Symbols denote LCSOs, solid lines denote regressions for individual datasets,
and dotted lines denote pooled regressions over all datasets.
150 -
Fathead Minnow
(Thurston et al. 1983)
O
o
Channel Catfish
(Gary 1976)
10
20
30
Rainbow Trout
(Thurston and Russo 1983) 100
80
10
10
20
30
30
20
15
8
6
4
Channel Catfish
(Colt & Tchobanoglous 1976) 40
30
- - - • . T u 'V 7n
15
10
8
I I I n
Rainbow Trout
- (Ministry of Tech 1968)
O
O
O
1 1 1
600 r
Atlantic Salmon
(Knopf 1992)
10
20
30
10
20
30
20
30
O)
o
m
O
.
O
o
40
30
20
15
10
Bluegill
(Roseboom & Richey 1977)
10
20
30
80
60
40
30
20
15
10
Channel Catfish 60
(Roseboom & Richey 1977)
40
30
10
8
10
20
30
Largemouth Bass
(Roseboom & Richey 1977)
10
20
Rainbow Trout 40
_ (Reinbold & Pescitelli 1982a) 3Q
20
15
10
30
10
20
30
Bluegill
(Reinbold & Pescitelli 1982a) 40
30
20
15
10
10
20
30
Johnny Darter A
(Nimmo et al.
10
20
30
60
40
30
20
15
10
Fathead Minnow
(Reinbold & Pescitelli 1982a) 100
80
10
20
30
Three-Spined Stickleback
Fathead Minnow
(DeGraeve et al. 1987)
20 30 0 10 20
TEMPERATURE (C)
400
300
60
Channel Catfish
(DeGraeve et al. 1987)
30
10
20
30
60
40
30
20
15
10
8
Fathead Minnow
(Nimmo et al. 1989)
10
20
30
13
-------
Figure 3. The effect of temperature on pH-adjusted acute ammonia toxicity in terms of
total ammonia. LCSOs are adjusted to the mean pH of the dataset based on the
pooled relationship of acute toxicity to pH. Symbols denote LCSOs, solid lines
denote regressions for individual datasets, and dotted lines denote pooled
regression over all datasets.
1 50
1 00
80
60
40
30
20
Fathead Minnow
(Thurston et al. 1983)
1 0 20
30
80
60
40
30
20
1 5
1 0
Rainbow Trout
(Thurston and Russo 1983) 100
80
60
40
30
20
—L 1 5
30 0
Channel Catfish
(Gary 1976)
1 0 20
30
20
1 5
8
6
4
Channel Catfish
(Colt & Tchobanoglous 1976) 40
30
1 5
1 0
8
Rainbow Trout
- (Ministry of Tech 1968)
- v o
—
1 1 1
600
400
300
200
1 50
1 00
80
Atlantic Salmon
(Knopf 1992)
1 0 20
30
1 0 20
30
0 10 20 30
1 0 20
30
05
£
O
m
O
40
30
20
1 5
1 0
8
6
Bluegill
(Roseboom & Richey 1977)
80
60
40
30
20
1 5
1 0
Channel Catfish 60
(Roseboom & Richey 1977)
40
30
20
1 5
1 0
LargemouthBass 40
(Roseboom & Richey 1977)
20
1 5
1 0
8
Rainbow Trout
_ (Reinbold & Pescitelli 1982a)
40
20
1 5
1 0
Bluegill
(Reinbold & Pescitelli 1982a) 40
30 -
10 20
30
10 20
30
1 0 20
30
10 20
30
1 0 20
30
60
FatheadMinnow
(Reinbold & Pescitelli 1982a) 100
80
Three-Spined Stickleback
Striped Bass
— (Hazel et al. 1971)
Fathead Minnow
(DeGraeve et al. 1987)
400
300
1 00
80
60
Channel Catfish
(DeGraeve et al. 1987)
10 20 30 0 10 20 30
TEMPERATURE (C)
10 20
30
60
40
30
20
1 5
1 0
Fathead Minnow
(Nimmo et al. 1989)
10 20 30
14
-------
the data are adjusted for the fact that pHs were lower in the low-temperature tests than
in the high-temperature tests. Slopes from regression analyses of datasets in Figure 3
range from -0.015 to 0.013, compared to a range from 0.015 to 0.054 when expressed
in terms of un-ionized ammonia (Erickson 1985). This narrower range of slopes in
terms of total ammonia nitrogen also argues for use of total ammonia, rather than un-
ionized ammonia, because there is less uncertainty associated with the generic
relationship. For datasets with just two points, Figures 2 and 3 also show the slopes for
comparative purposes. Based on the typical uncertainty of LCSOs, these slopes also
would not be expected to be significant, except perhaps for the Johnny darter data of
Nimmoetal. (1989).
A multiple least-squares linear regression (Draper and Smith 1981) using all datasets
(with a common slope for all datasets and separate intercept for each dataset) was
conducted, both with and without pH adjustment. The results of these pooled analyses
are plotted as dotted lines in Figures 2 and 3 to show that the residual errors for the
common regression line compared to the individual regression lines are not large
relative to the typical uncertainty of LCSOs. To better show the overall fit of the
common regression line, the data are also plotted together in Figure 4 by dividing each
point by the regression estimate of the LC50 at 20°C for its dataset. This normalization
is done strictly for data display purposes because it allows all of the datasets to be
overlaid without changing their temperature dependence, so that the overall scatter
around the common regression line can be better examined. The data show no
obvious trend, with the best-fit slope explaining only 1 % of the sum of squares around
the means for the pH-adjusted data and 0% for the unadjusted data. The one available
chronic dataset (DeGraeve et al. 1987) also shows no significant temperature effect
when expressed in terms of total ammonia nitrogen (Figure 5) and adjusted for pH
differences among the tests. (These tests and the calculation of the LC20s are
discussed in detail later.)
Based on the small magnitude and the variability of the effect of temperature on total
ammonia acute and chronic toxicity values for fish, the 1998 Update did not include
temperature as a modifying factor for a total ammonia criterion. For invertebrates, it
should be noted that the 1998 Update's assumption that temperature had no effect on
the toxicity of total ammonia differs from the 1984/1985 criteria document's assumption
that temperature has no effect on the toxicity of un-ionized ammonia. This
inconsistency is resolved during the 1999 re-examination of data, to be discussed
shortly, by incorporating a relationship between temperature and total ammonia toxicity
to invertebrates. That relationship, however, does not affect the CMC because
invertebrates are not among the acutely sensitive taxa.
The amount of uncertainty in this approach to the CMC can be demonstrated to be
small by considering how the criterion would differ if total ammonia toxicity was
adjusted based on the slopes in various datasets. Because the bulk of the toxicity data
used in the derivation of the criterion is within a few degrees of 20°C, the temperature
relationship used has very little effect on the criterion near this temperature, but rather
has the greatest effect on the criterion at much higher or lower temperatures. If the
15
-------
Figure 4. The effect of temperature on normalized acute ammonia toxicity in terms of total
ammonia. Data were normalized by dividing measured LCSOs by regression
estimates of LCSOs at 20°C for individual datasets for Figure 2 (top plot) and
Figure 3 (bottom plot).
2
1.5
1
0.8
0.6
0.4
2
1.5
1
0.8
0.6
0.4
10
15
Not pH Adjusted
CM
20 25 30
pH Adjusted
10 15 20
TEMPERATURE (C)
25
30
V ChannelCatfish
Rainbow Trout
© Bluegill
H ChannelCatfish
A LargemouthBass
^ Rainbow Trout
<$> Bluegill
• Fathead Minnow
• Striped Bass
A Stickleback
^ Fathead Minnow
+ ChannelCatfish
O AtlanticSalmon
© Johnny Darter
• Fathead Minnow
16
-------
Figure 5. The effect of temperature on chronic ammonia lethality to fathead minnows in
terms of total ammonia (DeGraeve et al. 1987). Symbols denote LC20s and 95%
confidence limits and lines denote linear regressions of log LC versus
temperature. Figure on left is for estimated LCSOs at test pH and figure on
right is for LCSOs adjusted to pH=7.5 based on pooled relationship of
chronic toxicity to pH.
100
50
D)
o
CM
o
TO
'c
o
E
"ro
-i— •
o
20
10
Data Not pH Adjusted
100 i-
50
20
10
Data Adjusted to pH 7.5
10 15 20 25
Temperature (C)
30
10 15 20 25
Temperature (C)
30
17
-------
average slope for the pH-adjusted acute data from Figure 4 is used, the total ammonia
CMC at 5°C would be only about 6% higher than at 20°C. The smallest and largest
slopes from the acute regressions for individual species in Figure 3 would produce a
range from 40% lower to 68% higher at 5°C than at 20°C, but this greatly overstates
the uncertainty because effects on a CMC derived from many datasets should not be
near these extremes.
1999 Re-examination of Temperature Dependence - Acute Toxicity
The previous section, reproduced with relatively few changes from the 1998 Update,
included an analysis of available data on the temperature dependence of acute
ammonia toxicity to fish. These data (in Figures 2, 3, and 4) consisted of 20 different
data sets drawn from 11 different studies and included nine different species, four of
these species being in more than one study. Data from Arthur et al. (1987) were not
used in the 1998 analysis because those authors reported concerns about factors
confounding temperature in their data set. Linear regression analysis of log LC50 (total
ammonia basis) versus temperature was conducted on each data set, both with and
without correcting for pH as a confounding factor. No consistent trend with temperature
was observed and only one data set showed a slope different than zero at the 0.05
level of statistical significance. Therefore, a pooled linear regression analysis was
conducted across all data to derive an average slope, which was very close to zero and
also not statistically significant. On the basis of this analysis, the 1998 Update did not
include any temperature dependence for criteria to protect fish from acute ammonia
toxicity.
In response to public comment (U.S. EPA, 1999), the 1998 analysis was re-examined.
This re-examination indicated that it is appropriate to handle the temperature
dependencies offish and invertebrates separately. For invertebrates, the inclusion of
the Arthur et al. (1987) data in the regression analysis, yields a change in the
temperature dependency that is ultimately reflected in the difference between the 1999
CCCandthe1998CCC.
In the 1998 Update, EPA did not use the Arthur et al. (1987) data because of those
authors concerns that other variable factors in their tests, conducted during different
seasons, might have had a potential to confound their results. In re-examining their
data in response to comments, however, EPA found that most of the fish data from
Arthur et al. showed behavior similar to that from numerous other investigators: that is,
little relationship with temperature when expressed as total ammonia. Consequently, it
was concluded that the other variable factors were unlikely to be confounding the
results.
For fish, although the temperature dependency is unchanged from 1998, additional
documentation is provided here, primarily because the apparent difference between
fish acute and chronic temperature dependencies is now used in the projection of the
invertebrate chronic temperature dependency.
18
-------
First presented here will be more details on the regression analyses of the individual
data sets conducted for the Update, plus similar analyses of the data of Arthur et al. A
linear regression was conducted on each data set using the equation:
log(LC50T) = log(LC5020) + S-(T-20) (5)
where LC50T is the total ammonia LC50 at temperature T, S is the slope of log LC50
versus temperature, and LC5020 is the estimated total ammonia LC50 at 20°C. For
completeness, this effort included data sets with just two points, although the
regression analysis then provides a perfect fit and has no residual error, so that
confidence limits, significance levels, etc. cannot be evaluated using normal methods.
In such cases, the mean squared error (MSE) of data around the regression was
assumed to be equal to the weighted mean residual MSE for the larger data sets, so
that approximate significance levels could be determined.
Fish acute data:
Table 1 presents the results of the regression analysis for each data set, with data
adjusted to pH=8 based on the average pH relationship used in the 1998 and 1999
Updates. Plots of these relationships (except for Arthur et al. 1987) are in Figure 3 in
the previous section.
Of the 24 entries in Table 1, nearly half (11) have very small slopes of between -0.006
and +0.006, a range which corresponds to a factor of 1.3 change or less in LC50 for a
20°C temperature change and is less than normal data variability. Of these 11, five
have positive slopes and six have negative slopes. Of the 13 entries with steeper
slopes, five have positive slopes and eight have negative slopes. Among the data sets
used in the Update, only two of the regressions are statistically significant at the 0.05
level, one with a negative slope and one with a positive slope, although two other sets
(for fathead minnows from DeGraeve et al. 1987 and Reinbold and Pescitelli 1982) are
close to being significant. (The level of significance for the Johnny darter data set
differs from what was reported in the previous section because it consists of two
different sets which were analyzed separately in the 1998 analysis, but combined here
because they were not significantly different.) Of the five data sets from Arthur et al.
(1987), only one is significant at the 0.05 level. For species with more than one entry,
slopes vary considerably. This general lack of statistical significance and consistency
precludes any reliable assessments based on these individual analyses.
The 1998 Update therefore conducted a pooled regression analysis to determine
whether the combined acute toxicity data sets indicated any significant average trends
with temperature. Table 2 summarizes the mean trends determined in various pooled
analyses. The first entry is the pooled analysis conducted for the 1998 Update, which
included all the data in Table 1 above except the fish data of Arthur et al. (1987). The
slope from this pooled analysis was very small (-0.0023), and was not statistically
significant despite the large number of data. The second entry adds the fish data of
19
-------
Table 1 . Results of regression analysis of logLCSO (mg/L total ammonia) versus temperature (°C) for
individual data sets on the temperature dependence of acute ammonia toxicity.
Reference/
Species
Thurston et al. (1983)
Fathead Minnow
Thurston and Russo (1 983)
Rainbow Trout
Gary (1976)
Channel Catfish
Colt and Tchobanoglous (1976)
Channel Catfish
Ministry of Technology (1967)
Rainbow Trout
Roseboom and Richey (1 977)
Bluegill Sunfish
Roseboom and Richey (1 977)
Channel Catfish
Roseboom and Richey (1 977)
Largemouth Bass
Reinbold and Pescitelli (1982)
Rainbow Trout
Reinbold and Pescitelli (1982)
Bluegill Sunfish
Reinbold and Pescitelli (1982)
Fathead Minnow
Hazel etal. (1971)
Striped Bass
Hazel etal. (1971)
Three-Spined Stickleback
DeGraeve etal. (1987)
Fathead Minnow
DeGraeve etal. (1987)
Channel Catfish
Knopf (1992)
Atlantic Salmon
Knopf (1992)
Atlantic Salmon
Nimmo etal. (1989)
Johnny Darter
Nimmo etal. (1989)
Fathead Minnow
Arthur etal. (1987)
Fathead Minnow
Arthur etal. (1987)
Rainbow Trout
Arthur etal. (1987)
Channel Catfish
Arthur etal. (1987)
White Sucker
Arthur etal. (1987)
Walleye
Slope
(95% CL)
-0.0014
(-0.013.+0.013)
+0.0059
(-0.017.+0.029)
+0.0028
(-0.008.+0.013)
+0.0004
(-0.037.+0.038)
+0.0008
(-0.018.+0.019)
+0.024
(-0.025.+0.073)
+0.020
(-0.029.+0.069)
-0.0029
(-0.040.+0.034)
-0.0088
(-0.028.+0.010)
-0.0004
(-0.027.+0.026)
-0.0153
(-0.031 ,+0.009)
-0.0163
(-0.057.+0.025)
-0.0106
(-0.053.+0.032)
-0.0052
(-0.012.+0.002)
-0.0088
(-0.016.-0.002)
-0.0035
(-0.027.+0.020)
+0.0163
(-0.075.+0.108)
+0.021
(+0.000.+0.043)
+0.0070
(-0.014.+0.028)
-0.032
(-0.059.-0.004)
-0.0100
(-0.053.+0.033)
-0.0058
(-0.038.+0.027)
+0.0007
(-0.23.+0.25)
-0.038
(-0.327.+0.250)
logLC5020
(95% CL)
1.641
(1.582,1.700)
1.350
(1.204,1.495)
1.676
(1.593,1.758)
1.604
(1.350,1.858)
1.231
(1.010,1.452)
1.089
(0.803,1.375)
1.482
(1.196,1.768)
1.237
(0.972,1.502)
1.396
(1.159,1.632)
1.370
(1.059,1.681)
1.429
(1.243,1.615)
1.274
(1.105,1.443)
1.390
(1.214,1.567)
1.617
(1.563,1.670)
1.648
(1.595,1.701)
1.715
(1 .406,2.025)
1.636
(0.405,2.866)
1.463
(1.248,1.678)
1.568
(1.353,1.782)
1.762
(1 .493,2.030)
1.348
(0.937,1.758)
1.558
(1.230,1.886)
1.902
(1.657,2.147)
1.216
(-1 .91 1 ,4.343)
Residual so
(r2)
0.112
(<1%)
0.121
(2%)
0.093
(2%)
0.016
(2%)
0.051
(1%)
-
-
-
0.088
(29%)
0.128
(0%)
0.076
(89%)
0.076
(60%)
0.081
(36%)
0.061
(36%)
0.061
(62%)
0.097
(7%)
0.054
(84%)
0.072
(90%)
-
0.105
(92%)
0.158
(16%)
0.030
(84%)
0.048
(1%)
0.306
(74%)
FREGR
(a)
0.06
(0.81)
0.30
(0.59)
0.32
(0.58)
0.02
(0.91)
0.03
(0.88)
0.95
(0.33)
0.68
(0.41)
0.02
(0.88)
1.63
(0.27)
0.00
(0.96)
16.6
(0.06)
2.93
(0.23)
1.14
(0.40)
3.33
(0.12)
9.76
(0.02)
0.22
(0.067)
5.18
(0.26)
18.1
(0.05)
0.42
(0.52)
24.8
(0.04)
0.56
(0.51)
5.15
(0.26)
0.01
(0.92)
2.84
(0.34)
20
-------
Arthur et al.; it does result in a statistically significant trend. The mean slope (-0.0058)
is still small, but does amount to a 23% decrease in LC50 per 20°C increase in
temperature. However, this slope is heavily influenced by two points with high residual
(>3o) deviations. One of these points is a test at 3.4°C by Arthur et al. (1987) with
fathead minnows, which showed much greater effects of low temperature than other
studies with the same species. The other point is for a test at 22.6°C by Arthur et al.
(1987) with walleye, which showed very high sensitivity and was part of a set of three
tests which used fish from different sources, potentially confounding the temperature
effects. Without these two data, the regression has an even lower slope and is not
significant at the 0.05 level (third entry in Table 2). Overall, these analyses of the fish
acute data suggest a weak overall trend of higher LCSOs at low temperatures, with a
logLCSO versus temperature slope in the -0.002 to -0.006 range, but of questionable
statistical significance.
It is also useful to consider separately the overall trends for different fish species.
Table 1 includes multiple studies with fathead minnows, rainbow trout, channel catfish,
and bluegill sunfish. Table 2 includes the results of pooled analyses for each of these
species, both with and without data from Arthur et al. (1987). For rainbow trout,
bluegill, and channel catfish, the regressions were not statistically significant. The
bluegill data indicated virtually no temperature effect, whereas weak trends similar to
the pooled analyses over all data sets were suggested in the channel catfish data
(slope = -0.0030 without and -0.0034 with Arthur et al. data) and rainbow trout data
Table 2. Results of regression analysis of log LC50 (mg/L total ammonia) versus temperature (°C)
for pooled data sets on the temperature dependence of acute ammonia toxicity to fish.
Data Sets Pooled
All Data excluding Arthur et al.
All Data including Arthur et al.
All Data including Arthur et al. except "Outliers"
Fathead Minnow excluding Arthur et al
Fathead Minnow including Arthur et al.
Fathead Minnow including Arthur et al. exc "Outlier"
Rainbow Trout excluding Arthur et al.
Rainbow Trout including Arthur et al.
Channel Catfish excluding Arthur et al.
Channel Catfish including Arthur et al.
Bluegill Sunfish
Slope
(95% CL)
-0.0023
(-0.0057.+0.0011)
-0.0058
(-0.0094.-0.0022)
-0.0030
(-0.0063.+0.0002))
-0.0063
(-0.0122.-0.0005)
-0.0105
(-0.0169.-0.0049)
-0.0073
(-0.0129.-0.0017)
-0.0013
(D.0122,+0.0096)
-0.0034
(-0.0133.+0.0064)
-0.0030
(-0.0091 ,+0.0031)
-0.0034
(-0.088.+0.021)
+0.0006
(-0.0172.+0.0184)
Residual so
(r2)
0.105
(2%)
0.122
(8%)
0.105
(3%)
0.106
(11%)
0.120
(25%)
0.106
(15%)
0.109
(<1%)
0.115
(2%)
0.088
(4%)
0.085
(6%)
0.120
(<1%)
FREGR
(a)
1.79
(0.18)
10.3
(O.01)
3.52
(0.06)
4.76
(0.04)
13.4
(O.01)
6.85
(0.01)
0.06
(0.80)
0.51
(0.48)
1.05
(0.32)
1.64
0.21)
0.01
(0.92)
21
-------
(slope = -0.0014 without and -0.0034 with Arthur et al. data). For fathead minnow, the
pooled analyses were statistically significant and stronger, with slopes ranging from -
0.0063 to -0.0105 depending on the treatment of data from Arthur et al. Such slopes
for fathead minnow would result in moderate effects over a broad temperature range: a
20°C decrease in temperature would result in a 33% to 62% increase in LC50.
However, this species is not sensitive enough that this would affect the acute criterion
values. For the species used in the acute criterion calculations, no temperature
correction for acute toxicity is appropriate due to the lack of any significant trend over
all data sets.
Invertebrate acute data:
Unlike fish, available acute toxicity data for invertebrates indicates that their acute
sensitivity to ammonia decreases substantially with decreasing temperature. The 1998
Update noted this temperature dependence, but did not present any analysis of it
because tested invertebrates were sufficiently tolerant to acute ammonia exposures
that this dependence would not affect the acute ammonia criterion. The 1998 Update
also noted that this temperature dependence should be a consideration in setting low
temperature chronic criterion, but did not provide any specific analysis regarding this.
This section will provide an analysis of available information on the temperature-
dependence of invertebrate acute ammonia toxicity, to be used later for estimating the
temperature-dependence of chronic ammonia toxicity.
Arthur et al. (1987) provide the only available data on the temperature dependence of
acute ammonia toxicity to invertebrates. As noted earlier, these toxicity tests did not
specifically test temperature effects, but rather were seasonal tests in which various
chemical characteristics of the tests water varied as well as temperature. Test
organisms were whatever were available in outdoor experimental streams at the time of
the test, so the size, life stage, and condition of the organisms also varied. The authors
of this study expressed some doubt as to how much of the effects they observed were
actually due to temperature. However, for invertebrates, they did observe strong
correlations of total ammonia toxicity with temperature. Confounding factors might
contribute somewhat to this correlation, but temperature is still likely the primary
underlying cause. If other factors were largely responsible for the apparent effects of
temperature, it would be expected that strong correlations with temperature would also
be evident in their fish data. However, as discussed above, the fish data usually
showed much weaker effects of temperature, similar to other studies with fewer
confounding factors.
These invertebrate acute data were analyzed using the same regression model and
techniques as discussed above for fish. The study of Arthur et al. (1987) included data
sets for nine invertebrate species, but two of these sets were not included in the
analysis because they consisted of two tests at temperatures only 3°C apart. For the
other species, the number of tests ranged from 2 to 6, with temperature ranges of from
9°C to 21 °C. Table 3 summarizes the regression results for the data sets of each
species and for pooled analyses conducted on (a) all seven species, and (b) three
22
-------
species that had more than two tests and a temperature range of at least 15°C. All
data were corrected to pH 8 based on the average acute pH relationship (described
later). All species show substantially greater tolerance to ammonia at lower
temperatures, and in most cases the significance level of the regression is better than
0.05. (As for the analysis of the fish data, when there were just two tests for a species,
the significance level for the individual analysis is based on the MSE from the pooled
analysis.) The slope of log LC50 versus temperature does not vary widely, ranging
from -0.028 to -0.046 and being -0.036 for both pooled analyses. Figure 6 provides
plots of this data and the regression lines comparable to those for fish previously
shown in Figures 3 and 4.
Again, because invertebrates are not among the species acutely sensitive to ammonia,
the invertebrate acute temperature slope does not affect the formulation of the acute
criterion. It will be used subsequently, however, in formulating the invertebrate chronic
temperature slope, which ultimately will affect the formulation of the chronic criterion.
1999 Re-examination of Temperature Dependence - Chronic Toxicity
Fish chronic data
As in the 1998 Update, the only available data on the temperature dependence of
chronic ammonia toxicity are from the study by DeGraeve et al. (1987) on survival of
juvenile fathead minnows during 30-day exposures to ammonia at temperatures
ranging from 6°C to 30°C. In contrast to acute toxicity, which for fathead minnows
showed sensitivity to be slightly reduced at low temperatures, this data on chronic
toxicity suggested greater sensitivity at low temperatures. However, this trend was
small, at least once the confounding effect of pH was corrected for, and not statistically
significant. Based on this analysis, the 1998 Update treated effect concentrations for
chronic ammonia toxicity to fish as it did for acute toxicity: as being invariant with
temperature. However, the 1998 Update also noted that, if seasonal variations in
temperature cause a shift in what endpoints the criterion should be based on, the
chronic criterion could have a seasonal temperature dependence even if effect
concentrations for specific chronic endpoints do not vary with temperature. (This is
discussed in a later section on seasonality.)
This section will provide more details regarding the analysis of the chronic toxicity data
from DeGraeve et al. (1987), and a comparison of its temperature dependence to that
of acute toxicity in the same study. Figure 5 showed the temperature dependence of
acute and chronic effect concentrations from this study.
An important issue in this analysis is the confounding effect of pH on the apparent
effect of temperature, because pH increased with decreasing temperature in these
chronic exposures. To examine what the effect of temperature is, the effect
23
-------
Figure 6.
Temperature Dependence of Acute Ammonia Toxicity
to Invertebrate Organisms from Arthur et al. (1987).
(Solid line denotes species slope; Dotted line denotes pooled slope)
500
200
100
I
Q.
"ro
ro
o
E
20
Physa gyrina
500
200 -
100 -
200 r
Musculium transversum
10
20
30
Helisoma trivolvis
200
o
O 100
O 50
H 20
O 0
Q
y 5000
<
2
cc
Q 2000
1000
500
20
1000
500
200
100
50 -
Asellus recovitzai
10
20 30
10 20
_L
30
2000
1000
500
200
100 -
Philarctus quaeris
10
20
30
200
Orconectes immunis
0 10
0
(N
8 5
iO/Estimated L(
K)
UJ
-1 1
T3
-------
concentrations should be adjusted to a common pH using an equation that accounts for
the effect of pH. A critical question then is what pH equation to use, because no study
exists for the effect of pH on this particular chronic endpoint (juvenile 30-day survival),
or on the interaction of pH and temperature effects. The 1998 Update used the pH
relationship derived for the chronic criterion. Of the pH relationships available, that one
is probably most appropriate, but entails some uncertainty. To evaluate how
conclusions about temperature effects will vary if the true pH relationship is different,
this analysis will also use the pH relationship for acute toxicity to fathead minnows from
Thurston et al. (1981). This relationship likely represents an extreme possibility; i.e., it
assumes that chronic toxicity pH relationships are the same acute ones, contrary to
what is indicated by available chronic studies.
Using the pooled chronic pH relationship (presented later in this document),
slope=0.010, significance=0.13, and r2=0.76. Using the fathead minnow acute pH
relationship, slope=0.0053, significance=0.32, and r2=0.45. In neither case is the
regression statistically significant at the 5 percent level, due to the amount and
variability of the data. Nevertheless, it should be noted that, in both cases, the chronic
data show an upward trend with temperature, in contrast to that observed for acute
toxicity. Even under the extreme assumption that these data have a pH relationship
similar to acute toxicity, the slope is 0.005, and is twice this under the assumption that
these data follow the chronic pH relationship. Thus, even if fathead minnows show
increased acute tolerance to ammonia at low temperatures, a similar assumption for
chronic toxicity is contraindicated.
The difference between acute and chronic temperature relationships can be better
assessed by looking at acute-chronic ratios. Figure 7 shows the temperature
dependence of the ratio of the acute LC50 to the chronic LC20. The chronic LC20s
used for the ACRs were normalized via the above two alternative pH relationships,
while the acute LCSOs were normalized only using the acute pH relationship. The
results show that for either pH normalization alternative, the ACRs are substantially
higher at lower temperatures than at higher temperatures. If the chronic data are pH-
normalized using the chronic pH relationship, the regression is significant at the 0.05
level, with a slope of -0.0155. If normalized using the acute pH relationship, the slope
is less (-0.0110), but even with this extreme assumption, there is only a 13 percent
probability that the regression slope arose by chance.
It is not surprising that acute-chronic ratios are higher at low temperature. Temperature
can affect toxicity in a variety of ways, one of which is simply to slow down responses.
This is evident in some reports on the effect of temperature on ammonia toxicity. For
example, for the rainbow trout data from Ministry of Technology (1967), there was little
effect of temperature on total ammonia LCSOs at 96 hours, but at shorter durations
LCSOs increased with decreasing temperature. The overall impact on the temperature
dependence of LCSOs and ACRs will depend on the duration of the acute toxicity test
and on the speed of action of acute ammonia toxicity in the species of concern.
However, temperature is likely to affect ammonia toxicity in multiple ways, some of
25
-------
Figure 7. Temperature-dependence of ammonia ACRs for fathead minnows.
(The choice of reference condition, pH=8 here versus pH=7.5 in Figure 5,
has no effect on slope or significance.)
o
CM
O
CD
"9
O
ro
o
Temperature-dependence of ammonia acute:chronic ratios for fathead minnows
10
o
in
O -s
i -j
CD
CD
M—
o
g
"co
a:
Chronic Data Adjusted to pH=8
Using Update Average
Chronic pH Relationship
Slope = -0.0155
r2 = 90%
Siglevel = 0.05
10 15 20 25
Temperature (C)
30
10
Chronic Data Adjusted to pH=8
Using Fathead Minnow
Acute pH Relationship
Slope = -0.0111
r2 = 80%
Siglevel = 0.11
10 15 20 25
Temperature (C)
30
which would alter acute and chronic toxicity similarly. Nonetheless, to some degree the
ratio of effect concentrations at different durations is expected to increase at lower
temperatures. This expectation, as well as the empirical evidence, argues against the
direct application of acute temperature relationships to chronic toxicity.
Invertebrate chronic projections
No data are available on the effect of temperature on chronic ammonia toxicity to
invertebrates. Because invertebrates are much more acutely tolerant at low
temperatures than at high temperatures, it is likely that their chronic toxicity would also
show some temperature dependence. However, as discussed above, there is reason
to expect acute and chronic toxicity to vary somewhat differently with temperature, with
acute-chronic ratios increasing at low temperature, especially for organisms for which
acute ammonia toxicity is not especially fast, which is the case for invertebrates
(Thurston et al. 1984). The observed trend in the fathead minnow ACRs provides
support for this expectation.
The critical question then becomes, how much of the acute temperature slope for
invertebrates should be assumed to apply to chronic toxicity? If this slope is
predominantly due to temperature-induced delays of acute toxicity, chronic toxicity
26
-------
might have very little slope. If this slope is not at all due to such delays, then all the
slope should be applied to chronic toxicity.
One option for an objective mathematical prediction of the invertebrate chronic slope is
to assume that the difference between acute and chronic slopes will be the same for
fish and invertebrates, potentially implying that the effect of temperature on the kinetics
of toxicity is roughly the same for fish and invertebrates. In this case the invertebrate
chronic slope would be the difference between -0.036, the average invertebrate acute
slope, and -0.016, the observed slope for fish acute-chronic ratios. This would yield an
invertebrate chronic slope of -0.020. This correction still applies most of the acute
slope to chronic toxicity, but recognizes that the chronic slope should probably be less
steep.
It is recognized that few data are available to define the Figure 7 fish ACR slope, and
that the assumption the invertebrate ACR slope would equal the fish ACR slope is quite
uncertain despite having some theoretical underpinning in the kinetics of toxicity.
Consequently a second option is to equate the invertebrate chronic slope to the
invertebrate acute slope (-0.036) minus one-half the fish ACR slope (-0.016/2). This
splits the difference between no correction and full correction for the fish ACR slope,
resulting in an invertebrate chronic slope of -0.028.
A third, related option is suggested from the appearance of data in the last two plots in
Figure 6, plots of "All Species" and "Physa, Crangonyx, Musculium". These plots
suggest a steeper invertebrate acute slope at higher temperatures than at very low
temperature. At greater than 10°C, these data also comfortably fit a slope of -0.044. If
such a slope were used to fit those data, however, a concentration plateau would need
to be imposed between 5 and 10°C to avoid over-estimating the acute effect
concentrations measured near 5°C. If the invertebrate chronic slope is obtained by
subtracting the full value of the fish ACR slope (-0.016), this would yield the same
invertebrate chronic slope, -0.028, as the option in the previous paragraph. In this
case, however, concentrations would be capped between 5 and 10°C in order to reflect
the implied attenuation of slope at low temperature relative to higher temperatures.
EPA selected this third option, a compromise between the first two options, for defining
the invertebrate chronic temperature slope in formulating the CCC, discussed later.
This provides a good fit to the available information.
27
-------
pH-Dependence of Ammonia Toxicity
The 1984/1985 ammonia criteria document identified pH as an important factor
affecting the toxicity of ammonia and used an empirical model to describe the pH-
dependence of ammonia toxicity when expressed in terms of un-ionized ammonia. The
major features of this empirical model were a slope for logLCSO versus pH which was
approximately 1 at low pH and decreased as pH increased until pH«8, above which the
slope was 0. Such a model closely mimics a joint toxicity model, which also has a
slope of 1 at low pH and a slope of 0 at high pH when ammonia toxicity is expressed in
terms of un-ionized ammonia. The empirical model was parameterized based on a
pooled analysis of four datasets concerning the effect of pH on the acute toxicity of
ammonia. This effect of pH was generally supported by several additional datasets
reviewed by Erickson (1985), although some variation among species was evident,
especially for channel catfish. A dataset concerning chronic ammonia toxicity
(Broderius et al. 1985) indicated a somewhat greater effect of pH than for acute toxicity
and was used as the principal basis for the pH-dependence of the CCC.
As explained in the overview of this update, the effect of pH on the toxicity of ammonia
will be described here largely in terms of the joint (combined) toxicity of un-ionized
ammonia and ammonium ion. However, there is some dispute about whether ammonia
toxicity merely involves such joint toxicity. Also, a variety of factors might affect the
combined toxicity of the two forms. Therefore, use of a simple, mechanistic joint toxicity
model is inadvisable, and the following "S-shaped" model will be used to describe the
pH dependence of total ammonia toxicity:
LIMH LIM,
LC50t = y + L-
' - -
where the subscript t denotes total ammonia, LIMH and LIML are asymptotic (limiting)
LCSOs at high and low pH respectively, and pHT is the transition pH at which the LC50
is the arithmetic average of LIMH and LIML. This model is justified by various data (see
the overview) and is consistent with joint toxicity of un-ionized ammonia and ammonium
ion. However, the model treats pHT as a fitted parameter, whereas if joint toxicity were
assumed it would be dictated by the pK of ammonia (see Equation 4) and the relative
toxicity of the two forms.
Use of LIMH and LIML as model parameters results in a simple equation, but is
inconvenient for data analysis for two reasons. First, when analyzing toxicological
variables across multiple datasets, an important issue is whether the shapes of the
curves are similar among the datasets. For making such comparisons and for
estimating the best average shape, it is necessary that each parameter of the equation
either is related only to the shape or is not related to the shape at all. For example, in
linear regression, the equation is generally expressed in terms of a slope and an
28
-------
intercept (i.e., the value of y at a specified value of x, such as x=0). The slope
completely defines the shape of the relationship, whereas the intercept anchors the
relationship at a particular point and has no effect on the shape. For the nonlinear
regression used here, there needs to be one, and only one, "intercept" parameter that
specifies the LC50 at a particular pH, independent of the shape, whereas the other
parameters must describe aspects of the shape and not affect the intercept. In the
above equation, LIMH and LIML are both "intercepts" (at high and low pH, respectively),
and they also in part dictate the shape of the curve because the shape partly depends
on the difference between the two intercepts. Thus, it is not possible to completely
separate the shape from the intercepts. To eliminate this problem, the equation was
reformulated so that LIML is the only intercept parameter. This was accomplished by
using the parameter R = LIMH/LIML, which, along with pHT, defines the shape of the
curve:
LC50t =
The second shortcoming of the use of LIMH and/or LIML is that they are LCSOs at
extreme pHs which are not observed and are largely hypothetical; it is preferable to
have an "intercept" parameter that lies in the range of the observed data. Therefore,
the equation was reformulated to use the LC50t at pH=8 (LC50t8) as the intercept
parameter instead of LIML. Switching from LIML to LC50t8 requires use of a term that is
the ratio between LC50t8 and LIML:
LC5CX =
LC50t8
R
+
1
R
1
pH-pHT
(8)
All three of the above model equations are equivalent, differing only in the way in which
the parameters are formulated.
Unfortunately, analyses based on any of these three model equations can be subject to
serious problems with some datasets, especially for estimation of LIMH or R. This is
because LC50t is generally much greater than LIMH even at the highest pH in most
datasets (pH=8 to 9), so that the approach to this asymptotic value is very uncertain.
However, the pH is usually sufficiently high that un-ionized ammonia, although only a
small fraction of total ammonia, dominates toxicity and provides information about LIMH
and R that is not apparent when only total ammonia is examined. To address this
problem, the formulation of the model was changed by splitting the equation into two
parts:
29
-------
LC50U =
LC50(
t,8
R
R
(9)
LC50, =
LC50
t8
R
(10)
where LC50U and LC50, are the LCSOs expressed in terms of un-ionized ammonia and
ammonium ion, respectively, and LC50U + LC50, = LC50t. This approach more strongly
emphasizes the notion of joint toxicity, but still is somewhat empirical because pHT is a
fitted parameter. Regression methods for multiple response variables (see Appendix 2)
were used to fit this model to the available datasets.
Acute datasets evaluated included those cited in the 1984/1985 ammonia criteria
document and Erickson (1985), as well as more recent studies by Sheehan and Lewis
(1986), Schubauer-Berigan et al. (1995), Ankley et al. (1995), and Johnson (1995).
1. Sheehan and Lewis (1986) investigated the pH-dependence of acute ammonia
toxicity to channel catfish. LCSOs expressed in terms of un-ionized ammonia
increased with increasing pH, but less so than reported in most studies, although
Tomasso et al. (1980) also reported little effect of pH>7 on un-ionized ammonia
toxicity to the channel catfish.
2. Schubauer-Berigan et al. (1995) evaluated the effect of pH on the toxicity of
ammonia to the oligochaete Lumbriculus variegatus and to larvae of the dipteran
Chironomus tentans. Both species exhibited increases in 10-day un-ionized
ammonia LCSOs with increasing pH, but the increase for C. tentans was somewhat
larger than those for other species for which data are available, whereas those of
L. variegatus were smaller. Such interspecies differences would be of concern in
the derivation of the criterion if they substantially altered relationships for sensitive
species; these particular species, however, are sufficiently resistant to ammonia
that the pH relationship used for them has no impact on the criterion.
3. Ankley et al. (1995) tested the effect of pH on the toxicity of ammonia to the
amphipod Hyalella azteca in waters of three different ionic compositions. In all
three waters, 96-hr LCSOs expressed in terms of un-ionized ammonia increased
with pH, but the amount of increase was greater in waters with low ion
concentrations. These waters differed with respect to a variety of ions, so it is
uncertain which constituent is responsible for the difference in the effect of pH,
although recent work by Borgmann and Borgmann (1997) suggests that the
concentration of sodium is a major factor. These results not only indicate some
effect of the ionic composition of the test water on ammonia toxicity, but also
suggest that this composition might differentially affect the relative toxicity of un-
ionized ammonia and ammonium ion. In the low ion concentration test water,
30
-------
H. azteca was one of the most sensitive species tested at low pH and
consequences for the criterion will be considered later.
4. Johnson (1995) investigated the effect of pH on the chronic toxicity of ammonia to
Ceriodaphnia dubia in test waters of three different ionic compositions. In all three
waters, LCSOs expressed in terms of un-ionized ammonia increased with increasing
pH, but, unlike Ankley et al. (1995), the pH dependence was greater in waters with
higher, rather than lower, hardness.
Acute total ammonia LCSOs versus pH are presented in Figure 8 for all studies
analyzed; for the study of Ankley et al. (1995) with H. azteca, the small, medium, and
large symbols denote low, medium, and high ion concentrations in test waters. All
analyses were conducted in terms of total ammonia nitrogen, either as reported by the
authors or as converted by us from the reported un-ionized ammonia LC50, pH, and
temperature using the speciation relationship of Emerson et al. (1975). All of the
datasets show a strong trend of total ammonia LCSOs decreasing with increasing pH,
except that of H. azteca at low ion concentrations. There are, however, differences
among the datasets in the magnitude and shape of the trend. Some datasets show an
approach to an asymptote at low pH whereas others do not. In addition, C. tentans and
H. azteca show lower slopes than other species. Nevertheless, it would be speculative
to assign different relationships to different taxa, especially because the same or
closely related species show some variation. Consequently, all of the datasets were
used to determine an average, generic shape for the pH dependence.
Regression analyses were conducted individually on each dataset, and on the pooled
datasets assuming that only LC50t8 varied among datasets. The pooled analysis
estimated pHT to be 7.204 (95% confidence limits = 7.111 and 7.297) and R to be
0.00704 (95% confidence limits = 0.00548 and 0.00904). The individual regression
results are plotted as solid lines and the pooled analysis as dotted lines in Figure 8.
The data points and the common regression line from the pooled analysis are also
plotted together in Figure 9 by dividing each point by the LC50t8 for its dataset (this
normalized plot allows a different, combined perspective of the overall scatter of data
from the shape of the generic relationship not possible in Figure 8). Except for the
datasets for L. variegatus and H. azteca at low ion concentrations, the deviation of data
from this generic relationship at pH>7 is rather small and consistent with the typical
uncertainty of LCSOs. At pH<7, however, some of the deviations are substantial; some
species, most notably channel catfish and L. variegatus, have higher than expected
total ammonia LCSOs, whereas others, such as Daphnia sp. and H. azteca have lower
than expected LCSOs. Fortunately, these species are generally sufficiently resistant
that more accurately describing their pH dependence is unimportant for deriving a
CMC. Despite the variation among species at low pH, this generic relationship is
appropriate for criteria derivation, because it provides significantly higher values at low
pH, but not higher than those for fish species that are relatively sensitive at low pH, a
suitably conservative assumption for sensitive species for which data do not exist at
low pH.
31
-------
Figure 8. The effect of pH on acute ammonia toxicity in terms of total ammonia. Symbols
denote LC50s, solid lines denote regressions for individual datasets, and
dotted lines denote pooled regression over all datasets.
Fathead Minnow
(Thurston et al. 1981b)
10
Co ho Salmon
(Robinson-W ilson & Seim 1975) 4
Daphnia sp.
(Tabata 1962)
Smallmouth Bass
(Broderius et al. 1985)
D)
O
in
O
Rainbow Trout
(Lloyd & Herbert 1960)
" Macrobrachium rosenber
(Armstrong et al. 1978)
400
200
100
Lumbriculus variegatus
(Schubauer-Berigan et a
Channel Catfish
(Sheehan & Lewis 1986)
C h ironomus tentans
(Schubauer-Berigan et al. 1995)
Hyalella azteca
(Ankley et al. 1995)
pH
32
-------
Figure 9. The effect of pH on normalized acute ammonia toxicity in terms of total
ammonia. Data were normalized by dividing measured LCSOs by regression
estimates of LCSOs at pH=8 for individual datasets from Figure 8. Data were
not normalized in any way for temperature.
o
m
O
.
O
o
Q
LU
N
o:
o
30 r
20
10
8
6
4
3
1
0.8
0.6
0.4
0.3
0.2
0.1
O Fathead Minnow
D Rainbow Trout
A Coho Salmon
V Daphnia sp.
<0> SmallmouthBass
O Green Sunfish
E] Rainbow Trout
A Prawn Larvae
V ChannelCatfish
O White Perch
• Guppy
• ChannelCatfish
A Lumbriculus
^ Chironomus
• Hyalella
PH
33
-------
For chronic toxicity, the data of Broderius et al. (1985) and Johnson (1995) were
analyzed in terms of total ammonia nitrogen using the same pH model (Figure 10). The
data used were EC25s reported by Johnson (1995) and EC20s calculated from the
data of Broderius et al. (1985) by regression analyses discussed later. (Because
Johnson's raw data were not available, EC20s could not be calculated, but the shape
of the curve should be the same for EC20s and EC25s.) Because the uncertainty of
the EC25s from Johnson (1995) was greater than that of Broderius et al. (1985) and to
prevent the greater number of data points for the invertebrate from overwhelming the
data for the fish, data points from Johnson (1995) were given a weighting factor of 0.5
in this analysis. These chronic data had a higher transition pH (7.688; 95% confidence
limits = 7.554 and 7.821) and a higher R (0.0232; 95% confidence limits = 0.0160 and
0.0334) than the acute data. The higher pHT is in accordance with differences
previously noted in the 1984/1985 criteria document regarding the pH dependence of
acute and chronic toxicity. Tests by Borgmann (1994) on the chronic toxicity of
ammonia to Hyalella azteca and by Armstrong et al. (1978) on the 6-day toxicity of
ammonia to Macrobrachium rosenbergii also support a lower slope for total ammonia
chronic toxicity versus pH at pH<8. The dependence of chronic ammonia toxicity on pH
appears to be sufficiently different from the dependence of acute ammonia toxicity to
justify use of two equations.
By substituting the values for R and pHT into Equation 8, the following equations are
obtained for describing the pH-dependence of acute values (AVs) and chronic values
(CVs) expressed in terms of total ammonia nitrogen:
AV, = (AV,8) ( ft0489 ^5 \
l '' 7-204-H H-7-204
CV, = (CVJ °'0676 * 2'91 (12)
\ t'°/ V -| + <|Q7.688-pH i + iQpH-7.688,/ v '
The range of the data used to derive these equations indicates that they should be
applicable from pH=6 to 9, although considerable error might exist at the lower end of
this range for certain species. Extrapolation below pH=6 is not advisable because of
the increasing scatter of the data from the common regression line at lower pH, and
extrapolation above pH=9 is not advisable because of inadequate knowledge about the
effect of the inhibition of ammonia excretion at high pH on results of toxicity tests
(Russoetal. 1988).
34
-------
Figure 10. The effect of pH on chronic ammonia toxicity in terms of total ammonia.
Symbols denote chronic effect concentrations and lines denote regressions of
effect concentrations versus pH. (For C. dubia, different symbols denote
different test water formulations. It is important to note the triangle, partially
obscured by the square at pH=6.5.)
Smallmouth Bass
(Broderius et al. 1985)
40
^ 20
O)
E,
cT 10
CM
O
LJJ
< 2
i
9
PH
100 i-
^ 40
O)
£
w 20
CM
O
LJJ
< 10
Ceriodaphnia dubia
(Johnson 1995)
P 2
1 U-
6
i
9
PH
35
-------
Formulation of the CMC
The scope of this project included a re-examination of the temperature and pH
relationships underlying the 1984/1985 Criterion Maximum Concentration (CMC).
Because the acute toxicity dataset contained in the 1984/1985 criteria document (U.S.
EPA 1985a) is relatively large, with tests involving species in 34 genera, the scope of
this project did not include a comprehensive literature search and critical review of all
of the acute toxicity data now available. Thus, the derivation here relies solely on
acute tests reported in Table 1 in the 1984/1985 criteria document, supplemented by
some newer studies relevant to the revised pH relationship.
However, some other newer studies of acute toxicity known by EPA were examined to
determine whether new data might materially affect the CMC. These studies include
Ankleyetal. 1995; Arthur et al. 1987; Bailey et al. 1985; Bergerhouse 1992,1993;
Dabrowska and Sikora 1986; DeGraeve et al. 1987; Diamond et al. 1993 (see
Appendix 1); Gersich and Hopkins 1986; Goudreau et al. 1993; Gulyas and Fleit 1990;
Hasan and Macintosh 1986; Henderson et al. 1961; Lee 1976; Mayes et al. 1986;
Monda et al. 1995; Nimmo et al. 1989; Russo et al. 1988; Sheehan and Lewis 1986;
Snell and Persoone 1989; Thomas et al. 1991; Tomasso and Carmichael 1986; Wade
1992; and Williams et al. 1986. These studies would add few new genera to the
dataset and their data are generally in the range already observed and would have little
impact on the four lowest Genus Mean Acute Values (GMAVs). The most significant
result of these studies is that some invertebrates are acutely sensitive to ammonia at
low pH and low ion concentration (Borgmann 1994; Ankley et al. 1995). Although new
data are not used in the derivation of the new CMC, they are compared to the new
CMC below.
All of the un-ionized ammonia acute values (LCSOs and ECSOs) in Table 1 of the
1984/1985 criteria document were converted to total ammonia nitrogen acute values,
using the reported temperatures and pHs and using the pK relationship from Emerson
et al. (1975). These total ammonia nitrogen acute values were then adjusted (see
Appendix 3) to pH=8 using the pH relationship developed above, with no adjustment for
temperature. These adjusted total ammonia nitrogen acute values (see Appendix 4)
were then averaged to determine Species Mean Acute Values (SMAVs) and GMAVs at
pH=8 using the procedure described in the 1985 Guidelines (U.S. EPA 1985b). These
SMAVs and GMAVs are presented in Table 4 (which is the same as Table 1 of the
1998 Update). Both are based on the test results of in Table 1 of the 1984/1985
criteria document. The GMAVs in the 1984/1985 and the 1998/1999 documents differ
because (a) pH and temperature are addressed differently in the two sets of
calculations, (b) the golden trout, cutthroat trout, and rainbow trout are now in a
different genus, and (c) and the new GMAVs are expressed in terms of total ammonia
nitrogen; the order of the genera is different mostly because the absence of an
invertebrate temperature normalization in either the 1984/1985 criteria document or the
1998 and 1999 Updates affects the 1984/1985 un-ionized LCSOs differently than the
1998 and 1999 total ammonia LCSOs. As noted earlier in the document, a temperature
36
-------
Table 4. Ranked Genus Mean Acute Values
Rank
34
33
32
31
30
29
28
27
26
25
24
23
22
21
Genus Mean
Acute Value
(ma N/l_a)
388.8
246.0
210.6
189.2
115.5
113.2
108.3
97.82
93.52
77.10
73.69
51.73
51.06
43.55
Species
Caddisfly,
Philarctus quaeris
Crayfish,
Orconectes immunis
Crayfish,
Orconectes nais
Isopod,
Asellus racovitzai
Mayfly,
Ephemerella grandis
Mayfly,
Callibaetis skokianus
Mayfly,
Callibaetis sp.
Beetle,
Stenelmis sexlineata
Amphipod,
Crangonyx pseudogracilis
Tubificid worm,
Tubifextubifex
Snail,
Helisoma trivolvis
Stonefly,
Arcynopteryx parallela
Snail,
Physa gyrina
Mottled sculpin,
Cottus bairdi
Mosquitofish,
Gambusia affinis
Fathead minnow,
Species Mean
Acute Value
(ma N/l_a)
388.8
1466.
41.27
210.6
189.2
175.6
75.93
113.2
108.3
97.82
93.52
77.10
73.69
51.73
51.06
43.55
Pimephales promelas
20 38.11 White sucker, 45.82
Catostomus commersoni
37
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Rank
19
18
17
16
15
14
13
12
11
10
Genus Mean
Acute Value
(ma N/l_a)
36.82
36.39
35.65
34.44
33.99
33.14
32.82
30.89
26.97
26.50
Species
Mountain sucker,
Catostomus platyrhynch
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia pulicaria
Brook trout,
Salvelinus fontinalis
Clam,
Musculium transversum
Channel catfish,
Ictalurus punctatus
Cladoceran,
Simocephalus vetulus
Guppy,
Poecilia reticulata
Flatworm,
Dendrocoelum lacteum
White perch,
Morone americana
Stoneroller,
Campostoma anomalurr
Smallmouth bass,
Species Mean
Acute Value
(mg N/La)
31.70
35.76
37.91
36.39
35.65
34.44
33.99
33.14
32.82
30.89
26.97
35.07
Micropterus dolomieu
Largemouth bass, 20.03
Micropterus salmoides
26.11 Walleye, 26.11
Stizostedion vitreum
25.78 Cladoceran, 25.78
Ceriodaphnia acanthina
25.60 Red shiner, 45.65
Notropis lutrensis
Spotfin shiner, 19.51
Notropis spilopterus
Steelcolor shiner, 18.83
Notropis whipplei
38
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a
Genus Mean Species Mean
Acute Value Acute Value
Rank (mg N/La) Species (mg N/La)
6 23.74 Brown trout, 23.74
Salmo trutta
5 23.61 Green sunfish, 30.27
Lepomis cyanellus
Pumpkinseed, 18.05
Lepomis gibbosus
Bluegill, 24.09
Lepomis macrochirus
4 21.95 Golden trout, 26.10
Oncorhynchus aquabonita
Cutthroat trout, 25.80
Oncorhynchus clarki
Pink salmon, 42.07
Oncorhynchus gorbuscha
Coho salmon, 20.26
Oncorhynchus kisutch
Rainbow trout, 11.23b
Oncorhynchus mykiss
Chinook salmon, 17.34
Oncorhynchus tshawytscha
3 17.96 Orangethroat darter, 17.96
Etheostoma spectabile
2 14.67 Golden shiner, 14.67
Notemigonus crysoleucas
1 12.11 Mountain whitefish, 12.11
Prosopium williamsoni
All values are total ammonia nitrogen at pH=8.
b Thurston and Russo (1983) conducted numerous acute toxicity tests with larval, juvenile, yearling,
and larger rainbow trout and demonstrated that large rainbow trout were measurably more sensitive
than other life stages. The average adjusted total ammonia nitrogen acute value for large rainbow
trout was 11.23 mg N/L. Therefore, this SMAV was lowered to 11.23 mg N/L in order to protect large
rainbow trout, as per the 1985 Guidelines (U.S. EPA 1985b).
39
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normalization for the invertebrates, although technically justifiable, is irrelevant to the
CMC formulation, due to the insensitivity of all invertebrate taxa to ammonia acute
toxicity.
The Final Acute Value (i.e., the fifth percentile) at pH=8 was calculated from this set of
adjusted total ammonia GMAVs to be 14.32 mg N/L. The SMAV for rainbow trout is
1 1 .23 mg N/L, and so the FAV is lowered to this value, as per the 1 985 Guidelines
(U.S. EPA 1985b), comparable to what was done in the 1984/1985 ammonia criteria
document. The CMC at pH=8 equals one-half of this FAV. Substitution of this CMC at
pH=8 for AVt8 in Equation 1 1 results in the following equation for expressing the CMC
as a function of pH:
PMr 0.275 39.0
CMC =
107.204-pH 1 + 1QPH -7.204
If the four genera (Oncorhynchus, Prosopium, Salmo, and Salvelinus) in the family
Salmonidae are excluded from the dataset in Table 4, the fifth percentile FAV with
salmonids absent is 16.8 mg N/L and the CMC is 8.4 mg N/L at pH=8; substitution into
Equation 1 1 gives the CMC as a function of pH:
riwir °-411 58-4
CMC =
107.204-pH 1 + 1QPH -7.204
Figure 11 shows the ranked GMAVs, the CMC with salmonids present, and the CMC
with salmonids absent, all at pH=8. The GMAVs represent LCSOs, whereas the CMCs
represent concentrations that are lethal to a minimal percentage of the individuals in
either the fifth percentile genus or a sensitive important species.
FAVs and CMCs are plotted in Figure 12, along with all of the individual total ammonia
acute values, unadjusted for pH or temperature, used in the calculations. The FAVs
show good correspondence with the lower range of the acute values. As discussed
above, more recent acute data are also in general accordance with the FAVs, except
that the Hyalella azteca LC50 from Ankley et al. (1995) at low ion concentration and
pH=6.5 is more than a factor of two below the FAV. Although some toxicity data are
expected to be below the FAV, inclusion of this genus in the calculation would have
resulted in a lower CMC, but only under these extreme water quality conditions and
only if the effects of both pH and ionic composition were described for each individual
genus, which is not possible with the data that are currently available.
40
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Figure 11. Ranked Genus Mean Acute Values (GMAVs) with Criterion
Maximum Concentrations (CMCs).
1000
100 ••
Q.
-i— •
CO
D)
CD
10
CMC salmonids absent
CMC salmonids present
Ranked Genera
41
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Figure 12. Acute LCSOs used in criteria derivation in
relationship to Final Acute Values (FAVs) and
Criterion Maximum Concentrations (CMCs).
1000
100 -
O)
E
.
O
O
10 -
O
FAV - 5th Pet
w/o Salmonids
rAV-Adult
Rainbow Trout
New CMC
w/o Salmonids
New CMC
w/ Salmonids
Old CMC
(10, 20 C)
Other Invertebrates
Salmonid Fishes
Nonsalmonid Fishes
~
7
\
8
\
9
PH
42
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Review and Analysis of Chronic Data
Due to the magnitudes of the acute-chronic ratios (ACRs) for ammonia, the ammonia
CCC is sufficiently low relative to the CMC that the CCC generally will be the
determining factor for permit limits. In the 1984/1985 ammonia criteria document, the
CCC is more uncertain than the CMC because (1) the CCC was calculated by dividing
the FAV by an ACR (thus including the uncertainties of both the FAV and the ACR) and
(2) fewer acceptable chronic toxicity tests were available and not all of them could be
used to derive ACRs. Additionally, depending on how they were derived, the individual
chronic values could differ with respect to the nature and degree of the toxic effects
they represented. To reduce this variability, EPA reviewed and analyzed all of the
chronic data used in the 1984/1985 criteria document and newer chronic data known to
the authors or suggested by peer reviewers to produce a more extensive and
consistent set of Chronic Values (CVs) that could be used to directly calculate a CCC
rather than to calculate it using ACRs. This procedure also has some limitations
because (a) the criterion usually decreases as the number of genera used in the
calculation of the 95th percentile decreases and (2) chronic tests have been conducted
with a larger proportion of the species that are acutely sensitive to ammonia than those
that are acutely resistant to ammonia.
The first two parts of this section describe how the chronic tests on ammonia were
reviewed and how the CVs were calculated. The third part discusses each chronic test
of which EPA was aware and presents the relevant results.
Review of Chronic Data
Each chronic dataset was subjected to the following two-step review process. The first
step was to determine whether the test methodology was acceptable for providing
information about a CV. A test was considered acceptable if the dilution water, control
mortality, experimental design, loading, etc., were consistent with ASTM Standards
E1193, E1241, and E1295 (ASTM 1997a,b,c). The concentration of dissolved oxygen
was also reviewed on the basis of the U.S. EPA (1986) dissolved oxygen criteria
document.
Reviewing the concentration of dissolved oxygen (DO) was difficult because (a) ASTM
Standards E1193, E1241, and E1295 (ASTM 1997a,b,c) express limits on high and low
concentrations of DO in terms of percent saturation, whereas U.S. EPA (1986)
expresses limits on low concentrations of DO in terms of the concentration itself, and
(b) neither specifies the limits in a way that can be used directly to interpret the kinds of
information that are given in most reports of the results of toxicity tests. Therefore, the
following rationale was used. The mean DO concentration needs to be within an
acceptable range, but limits expressed as long-term averages can allow excessively
low or high concentrations for too long a period. Conversely, a limit that must be
satisfied at all times can unnecessarily penalize investigators who make more than the
minimum number of measurements and ignores the fact that organisms can tolerate
43
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extreme concentrations for brief periods of time. Therefore, limits were placed on the
mean and the fifth and ninety-fifth percentiles of the DO concentrations. Use of limits
that are expressed in terms of the mean and the fifth and ninety-fifth percentiles is
straightforward when the mean and standard deviation are reported or when all of the
individual measurements are reported, but not when only the range is reported. If the
measured concentration of DO during a chronic test was reported as a range, the
lowest and highest values were considered to be concentrations that existed for at least
5 percent of the time during the test.
The limits used were:
1. A chronic test was considered questionable if either (a) the mean DO concentration
was below 60 or above 100 percent of saturation or (b) the concentration of DO
was below 50 or above 105 percent of saturation more than 5 percent of the time
during the test. These limits are similar to, but different from, the limits given in
ASTM Standards E1193, E1241, and E1295 (ASTM 1997a,b,c).
It is clear that 60 percent of saturation is the desirable lower limit in Section 11.2.1
of ASTM Standard E729 (ASTM 1997d); for practical reasons, this section allows
the concentration of DO to be between 40 and 60 percent of saturation during the
last 48 hours of 96-hr static acute tests. Because test organisms and BOD utilize
oxygen, when the concentration of DO is above 100 percent of saturation, it is quite
possible that the concentration of dissolved nitrogen is even more supersaturated,
which increases the possibility of gas bubble disease.
2. A chronic test was considered questionable if either (a) the mean measured DO
concentration was below the mean given below or (b) the DO concentration was
below the lower limit given below for more than 5 percent of the time during the
test:
Mean (mg/L) Lower Limit (mg/L)
Salmonids: 6.5 5.0
Warmwater fishes
Early life stages 6.0 5.0
Other life stages 5.5 4.0
Invertebrates 6.0 5.0
The first three means are presented on page 34 of U.S. EPA (1986) and are 0.5
mg/L above the concentrations given for "slight production impairment" on page 31.
U.S. EPA (1986) does not give a "mean" for invertebrates on page 34 and so the
last mean given above is 1 mg/L higher than the concentration given for "some
production impairment" on page 31. The lower limits are concentrations given on
page 31 for "moderate production impairment" or "some production impairment".
Regardless of how limits on the DO concentration are expressed, it is sometimes
difficult to apply them to the information that is reported concerning toxicity tests.
If there was no reason to believe that the test methodology was unacceptable, the
second step of the review process was to determine whether the test satisfied one of
the definitions given in the 1985 Guidelines for life-cycle, partial life-cycle, and early
44
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life-stage test. By definition, life-cycle tests can be conducted with either a fish species
or an invertebrate species, but partial life-cycle and early life-stage tests can only be
conducted with a fish species. The considerations that excluded the most tests were
that (a) tests that did not include the newly hatched life stage cannot be acceptable life-
cycle, partial life-cycle, or early life-stage tests, and (b) tests that did not study
reproduction cannot be acceptable life-cycle or partial life-cycle tests. Each test that
satisfied one of the definitions could provide one of three kinds of information:
1. If all of the tested concentrations of the toxicant were so high that all of them
caused unacceptable effects, the test will probably provide an upper limit on a CV,
i.e., the CV will be lower than the lowest tested concentration.
2. If all of the tested concentrations were so low that none of them caused an
unacceptable effect, the test will probably provide a lower limit on a CV, i.e., the CV
will be higher than the highest tested concentration.
3. If the low tested concentrations did not cause unacceptable effects but the high
tested concentrations did, the test will probably provide a CV.
If the test did not satisfy the requirements for any of the three kinds of tests, it was
necessary to determine whether the toxicant caused an unacceptable reduction in (a)
survival, reproduction, and/or hatchability over any period of at least seven days, or (b)
growth over a period of at least 90 days. If it caused either kind of unacceptable
reduction, the test will probably provide an upper limit on a CV or it might lower a CV
from an early life-stage test. If it did not cause either kind of unacceptable reduction,
the test cannot provide a CV or an upper or lower limit on a CV, but the test might
provide other useful information. Because the test is not an acceptable life-cycle,
partial life-cycle, or early life-stage test, an upper limit on a CV can be based on a
reduction in survival, reproduction, and/or hatchability over any period of at least seven
days, but it cannot be based on a reduction in weight gain for fewer than 90 days
because such a reduction might be temporary; such a test cannot provide a lower limit
on a CV because some other life stage might be more sensitive. Although some CVs
were based on histopathological effects in the 1984/1985 ammonia criteria document,
this current effort could find no justification for equating histopathological effects with
effects on survival, growth, and reproduction (see Appendix 5).
Calculation of Chronic Values
Chronic values used in aquatic life criteria documents have traditionally been based on
analysis of data to determine the highest tested concentration at which no relevant
toxicological variable had a value that was statistically significantly different from the
value for the control treatment (highest no observed adverse effect concentration,
HNOAEC) and the lowest concentration at which the value for at least one of the
relevant toxicological variables was significantly different from the value for the control
treatment (lowest observed adverse effect concentration, or LOAEC). When endpoints
are defined on the basis of such hypothesis testing of each tested concentration
against the control treatment, the CV is set equal to the geometric mean of the
HNOAEC and the LOAEC. Such a procedure has the disadvantage of resulting in
marked differences between the magnitudes of the effects corresponding to the
45
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individual CVs, due to variation in the power of the statistical tests used, the
concentrations tested, and the size and variability of the samples used (Stephan and
Rogers 1985). For example, the CVs reported in the 1984/1985 ammonia criteria
document corresponded to reductions from the control treatment of just a few percent to
more than fifty percent.
To make CVs reflect a uniform level of effect, regression analysis was used here both
to demonstrate that a significant concentration-effect relationship was present and to
estimate CVs with a consistent level of effect. Use of regression analysis is provided
for on page 39 of the 1985 Guidelines (U.S. EPA 1985b). The most precise estimates
of effect concentrations can generally be made for 50 percent reduction (EC50);
however, such a major reduction is not necessarily consistent with criteria providing
adequate protection. In contrast, a concentration that caused a low level of reduction,
such as an ECS or EC10, is rarely statistically significantly different from the control
treatment. As a compromise, the EC20 is used here as representing a low level of
effect that is generally significantly different from the control treatment across the useful
chronic datasets that are available for ammonia.
Regression analysis was performed on a chronic dataset only if the dataset met the
following conditions: (1) it contained a control treatment to anchor the curve at the low
end, (2) it contained at least four concentrations of ammonia to provide at least two
error degrees of freedom when the three-parameter equation is fit to a set of data, (3)
the highest tested concentration of ammonia caused >50 percent reduction relative to
the control treatment to anchor the curve at the high end, and (4) at least one tested
concentration of ammonia caused <20 percent reduction relative to the control
treatment to ensure that the EC20 was bracketed by tested concentrations of ammonia.
For life-cycle and partial life-cycle tests, the toxicological variables used in these
regression analyses were survival, embryo production, and embryo hatchability. For
early life-stage tests, the variables used were embryo hatchability, fry survival, and fry
growth; if ammonia apparently reduced both survival and growth, the product of these
variables (biomass) was analyzed, rather than analyzing them separately. For other
acceptable chronic tests, the toxicological variable analyzed was survival, reproduction,
hatchability, and/or growth as appropriate, based on the requirements stated above
concerning acceptability of chronic tests.
The regression model used was based on the logistic equation:
T =
This equation produces an "S-shaped" curve, with the toxicological variable of interest
(T) being at a control value (T0) at low concentrations, zero at high concentrations, and
declining at intermediate concentrations; the location and steepness of this decline are
determined by the parameters A and B, respectively. It is not argued that this equation
46
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embodies a mechanistic description of chronic toxicity, but rather that this is a useful
equation that incorporates the major features commonly observed in concentration-
effect relationships. Application of various forms and extensions of this equation to
toxicological data have been discussed by various authors, most recently by Moore and
Caux(1997).
To make the equation more directly interpretable with respect to effect concentrations
and to assist in determining confidence limits for such effect concentrations as the
EC20, the equation was reformulated to:
iQB(logC-logECp)|
where log ECp (i.e., the logarithm of the concentration causing T to be reduced by p
percent from T0) is a parameter rather than A. This equation was applied to each
dataset using nonlinear least-squares regression analysis (Draper and Smith 1981),
with p=20%. Software used for determining the least-squares solution was written in
FORTRAN using nonlinear search routines based on the Newton-Raphson method
(Dahlquist and Bjorck 1974).
Either transformation or weighting was applied to each dataset to improve the
homogeneity of the variance:
1 . When T was a percentage, the regression analysis was conducted on a
transformation T- of each data point T, as follows (Draper and Smith 1981):
V = arcsin^T/100) (17)
The regression equation was similarly transformed and the parameter T0 was
formulated to be the transformed effect.
2. When T was count data, the regression analysis was conducted on the square root
transformation of T, and the regression equation was similarly transformed (Draper
and Smith 1981).
3. When T was weight or biomass, no transformation was used, but each datum was
weighted by the inverse of its variance (Draper and Smith 1981). For weight data,
these weighting factors were based on standard errors (SEs) or standard
deviations (SDs) divided by IS|1/2 as reported by the authors. For biomass [B =
product of proportion survival (P) and weight (W) in early life-stage tests], the
variance was estimated as follows:
VAR(B) « W2-SEP2 + P2-SEw (18)
where SEP is the SE of P as reported by the authors or calculated as (P(1-P)/N)1/2,
and SEW is the SE of W as reported by the authors or calculated from their data.
47
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In addition to the dataset-specific transformation or weighting described above, all
regression analyses used a general weighting scheme to make the analyses more
appropriate for calculating EC20s. When this type of regression analysis is used to
calculate such low-effect concentrations as an EC20, lack of fit of the model at high-
effect concentrations can perturb the fit of the model at low-effect concentrations. If the
form of the regression equation is known to be completely accurate, such perturbation
is appropriate; in this case, however, the equation is not expected to describe the exact
form of the concentration-effect curve over the whole range of T. Because high effect
concentrations contain useful information about the nature of the curve, they should not
be excluded, but they should not be allowed to unduly influence the fit in the range from
0 to 50 percent reduction. Consequently, normal weights were given to data points up
to the first concentration with a 50% or greater reduction relative to the control
treatment and points at higher concentrations were weighted by half. An alternative
was to use a more complicated form of the logistic equation (e.g., Moore and Caux
1997), but such equations introduce their own uncertainties, especially for small
datasets, and their main effect on calculation of the EC20 is to reduce the influence of
data points at high effects, with much the same results as the weighting scheme used
here.
SEs of the regression parameters were calculated based on the variance/covariance
matrix of the linearized model at the least-squares solution (Draper and Smith 1981)
and 95% confidence limits for the parameters were calculated by multiplying these SEs
by the applicable t-statistic. Simulations showed that this procedure produces
confidence levels that are near or greater than 95%. The EC20 and its confidence
limits were computed by taking the antilog of the calculated logEC20 and its confidence
limits. Confidence limits on effect concentrations for percentages other than 20 and on
values for T at concentrations other than 0 were estimated by reformulating the
regression equation to use these values rather than EC20 and T0 as parameters, and
then recomputing the variance/covariance matrix at the least-squares solution to
determine the SEs of the new parameters.
Evaluation of the Chronic Data Available for Each Species
The following presents a species-by-species discussion of each chronic test on
ammonia evaluated by this project. For each species, the available chronic tests are
discussed in the following order: life-cycle tests, partial life-cycle tests, early life-stage
tests, other laboratory tests, and then results from a field study. Also presented are the
results of regression analysis of each dataset that was from an acceptable chronic test
and contained sufficient acceptable data. For each such dataset, Appendix 6 contains
a figure that presents the data and regression line, and a table with the regression
parameters. All analyses were conducted in terms of total ammonia nitrogen, either as
reported by the authors or as converted by us from the reported values for un-ionized
ammonia, pH, and temperature using the speciation relationship of Emerson et al.
(1975). When an EC20 could be determined, it is first reported as calculated by
regression analysis of the data at the pH and temperature of the test. Then, to facilitate
comparisons of sensitivities within and between species, each EC20 is adjusted to
48
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pH=8 using Equation 12, and adjusted to a temperature of 25°C using Equation 5 with
a slope of -0.028, per the discussions of the temperature- and pH dependency
sections. Species Mean Chronic Values (SMCVs) were derived when justified by the
data, and then Genus Mean Chronic Values (GMCVs) were derived when justified by
the SMCVs. All of the EC20s, SMCVs, and GMCVs that were derived are tabulated in
Table 5, which is located at the end of this section.
Musculium transversum (Sphaerium transversum) (Fingernail clam)
Anderson et al. (1978) conducted two 42-day tests of the effect of ammonia on
survival of field-collected juvenile clams whose length averaged 2.2 mm. The
results of the two tests were so similar that the data were pooled for analysis. The
lowest mean measured DO concentration in any treatment was 6.5 mg/L (77
percent of saturation) and the lowest individual measured concentration was 5
mg/L (60 percent of saturation). Survival in the control treatment and low ammonia
concentrations (<5.1 mg N/L) ranged from 79 to 90%, but decreased to zero at 18
mg N/L. Regression analysis of the survival data using an arcsine transformation
resulted in a calculated EC20 of 5.82 mg N/L at 23.5°C and pH=8.15. The EC20 is
6.63 mg N/L when adjusted to pH=8 and 25°C.
Sparks and Sandusky (1981) conducted a test similar to that of Anderson et al.
(1978) with field-collected juvenile clams whose average length was 2.1 mm.
Although this test used a better food, the test was conducted in the same laboratory
and used test organisms from the same pool in the Mississippi River as Anderson
et al. (1978); Sparks participated in both studies. The lowest mean measured DO
concentration in any treatment was 6.4 mg/L (73 percent of saturation) and the
lowest individual measured concentration was 5.0 mg/L (57 percent of saturation).
Survival in the control treatment was 92% and decreased with increasing
concentration of ammonia to 17% at 18 mg N/L. Effects on survival were evident at
lower concentrations, resulting in an EC20 of 1.23 mg N/L at 21.8°C and pH=7.80.
The EC20 adjusted to pH=8 and 25°C is 0.77 mg N/L. Although this EC20 is
substantially lower than that obtained by Anderson et al. (1978), the difference is
less than a factor of 10.
Zischke and Arthur (1987) studied fingernail clam growth, survival, and
reproduction in enclosures placed in experimental streams for periods of 4 to 10
weeks during a 16-month field study of the effects of ammonia (Hermanutz et al.
1987). Experiments during the first year showed reductions in survival of clams in
a stream in which the concentration of total ammonia nitrogen was approximately 2
mg N/L during the test period (Hermanutz et al. 1987), but not in a stream in which
the concentration was 0.7 mg N/L. The daily mean stream temperature ranged
from 20 to 25°C and pH ranged from 7.4 to 7.8 during this test period. During the
second year of the study, substantial effects occurred on reproduction of clams at 1
mg N/L (the lowest tested concentration of ammonia) at 24 to 26°C and pH=7.8 to
8.2 during the test period. Adjusted to pH=8, both years showed effects at about 1
mg N/L. These results are not included in Table 5 because results of field tests are
not used in the derivation of Final Chronic Values (U.S. EPA 1985b).
49
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The SMCV at pH=8 and 25°C is <2.26 mg N/L. This concentration is the geometric
mean of the adjusted EC20s for the two laboratory studies and is an upper limit on
the SMCV because the EC20s are based on survival of juveniles, which might not
be as sensitive to ammonia toxicity as early life stages. This SMCV is uncertain
due to the difference between the results of the two chronic tests. However, the
experimental stream data suggest that the SMCV should be close to 1 mg N/L. The
GMCV is also <2.26 mg N/L.
Ceriodaphnia acanthina
Mount (1982) conducted a life-cycle test that started with <1-day-old organisms and
proceeded until most of the control organisms produced three broods. The DO
concentration ranged from 5.7 to 6.4 mg/L (68 to 77 percent of saturation). Total
offspring production per treatment was unaffected at concentrations <21 mg N/L,
but reproduction was virtually absent at concentrations >77 mg N/L. Regression
analysis using a square root transformation resulted in an EC20 of 44.9 mg N/L at
pH=7.15 and 24.5°C. The EC20 adjusted to pH=8 and 25°C is 19.1 mg N/L, which
is the SMCV.
Ceriodaphnia dubla
Willingham (1987) conducted a 7-day life-cycle test starting with <1-day-old
organisms. The lowest mean measured DO concentration in any treatment was
6.04 mg/L (74 percent of saturation) and the lowest calculated fifth percentile of the
DO concentrations was 5.62 mg/L (69 percent of saturation). Production of young
during the third brood was unaffected at concentrations up to 2.8 mg N/L, but was
reduced at higher concentrations and was absent at 43 mg N/L. The EC20
calculated using regression analysis was 5.80 mg N/L at pH=8.57 and 26.0°C.
Adjusted to pH=8 and 25°C, the EC20 is 15.6 mg N/L.
Nimmo et al. (1989) conducted a 7-day life-cycle test at 25°C and pH=7.8 in water
from the St. Vrain River. The DO concentration was reported to be low in some
other tests that were conducted during this study, but it was not reported to be low
in this test. Based on the average number of neonates per original female, the
EC20 calculated using regression analysis and a square root transformation was
15.2 mg N/L. Adjusted to pH=8 and 25°C, the EC20 is 11.6 mg N/L.
As stated above in the discussion of the effect of pH on the toxicity of ammonia,
Johnson (1995) conducted twelve chronic tests on ammonia with C. dubia at four
pHs and three hardnesses. The lowest reported mean concentration of DO was
6.9 mg/L (82 percent of saturation). When adjusted to pH=8, the mean EC25s are
9.03, 7.46, and 17.1 mg N/L at average hardnesses of 42, 86, and 170 mg/L,
respectively. These mean adjusted EC25s are similar to the adjusted EC20s
obtained by Willingham (1987) and Nimmo et al. (1989). These EC25s are not
included in Table 5 because they are not EC20s and were calculated using a
different regression-type approach.
50
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Adjusted to pH=8 and 25°C, the two EC20s for C. dubia are 15.6 and 11.6 mg N/L,
which gives a SMCV of 13.5 mg N/L. For C. acanthina at pH=8 and 25°C, the
SMCV is 19.1 mg N/L, which gives a GMCV of 16.1 mg N/L.
Daphnia magna
Gersich et al. (1985) and Gersich and Hopkins (1986) reported results of a life-
cycle test that was conducted in water from the Tittabawassee River. This water
was probably an acceptable dilution water because it was apparently collected
upstream of all known point discharges (Alexander et al. 1986; James Grant,
Michigan Department of Environmental Quality, personal communication). The
lowest and highest measured DO concentrations were 8.8 and 9.2 mg/L (96 and
101 percent of saturation). No significant effects were found at concentrations up
to 4.2 mg N/L at pH=8.45 and 19.8°C, but progressively larger reductions were
found at concentrations of 9 to 36 mg N/L. The EC20 calculated from regression
analysis was 7.37 mg N/L.
In another life-cycle test, Reinbold and Pescitelli (1982a) found little reduction in
reproduction at 20 mg N/L, but a large reduction at 33 mg N/L. The measured DO
concentrations averaged 88 to 91 percent of saturation. The EC20 is 21.7 mg N/L
atpH=7.92and20.1°C.
Gulyas and Fleit (1990) conducted a 9-day chronic test to study the effect of
ammonia on development and growth. Concentrations that caused more than fifty
percent reduction compared to the controls were considered toxic. The "no effect
level" was reported to be 0.1 mg/L. No results from this test are included in Table 5
because neither survival nor reproduction was studied.
Adjusted to pH=8 and 25°C, the respective EC20s are 10.8 and 14.1 mg N/L. The
SMCV for this species is 12.3 mg N/L, which is the geometric mean of the two
adjusted EC20s; this is also the GMCV.
Crangonyx spp. (amphipod)
The available data for this species are not used for the reason(s) given in
Appendix 1.
Hyalella azteca (amphipod)
Borgmann (1994) conducted three tests that began with <1-week-old organisms, all
of which utilized weekly renewals and dechlorinated tap water originating from Lake
Ontario. One of the three tests lasted four weeks, but the other two lasted ten
weeks and produced data concerning both survival and reproduction. The results
of these last two tests were sufficiently similar that the results were analyzed
together. No information was reported concerning the DO concentration. Sufficient
raw data were obtained from the author so that each test chamber could be plotted
as a separate point for the combined regression analysis. Survival over the ten
weeks in the control treatment averaged 66.3 percent and reproduction per
chamber averaged 48 offspring. The 33.7% mortality in the control treatment is
51
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considered acceptable in a 10-week test because ASTM Standard E1706 (ASTM
1997e, Tables 10, 11, and 15) allows 10% mortality of H. azteca in a 4-day test and
allows 20% mortality in a 10-day test. In addition, although ASTM Standard E1706
allows 20% mortality in a 10-day test, Table 3 in Borgmann (1994) indicates that
only 11.6% of the controls died in four weeks.
At the lowest tested concentration, survival was reduced 25 percent relative to the
control treatment and reproduction was reduced 55 percent. Regression analysis
produced an EC20 of 0.88 mg N/L based on reproduction, but this EC20 is below
the lowest tested concentration because the dataset does not contain a
concentration that caused <20 percent reduction relative to the control treatment.
However, the confidence limits on the regression analysis indicate that the 55
percent reduction in reproduction caused by the lowest tested concentration is
statistically significant. Based on the raw data, the concentration of ammonia in the
lowest tested concentration was 1.58 mg N/L and the mean pH of this treatment
was 7.94. Therefore, the EC20 is <1.58 mg N/L at pH=7.94 and 25°C. Adjusted to
pH=8 and 25°C, the EC20 is <1.45 mg N/L. Even though chronic survival
appeared to be less sensitive than reproduction in this test, slightly more than 20%
mortality occurred at the lowest tested concentration; therefore, the LC50 for
chronic survival is <«1.45 mg N/L.
Because the test solutions were renewed once a week, the pH dropped and the
concentration of total ammonia increased between renewals; the average of the
weekly measured initial and final values was used for both pH and total ammonia.
The pH measured at the end of each week averaged 0.54 lower than the pH
measured at the beginning of each week in the control test chambers, and
averaged 0.78 lower in the two test chambers at the lowest tested concentration of
ammonia. Even though the average pH drop in the control test chambers for the
second test was 0.21 and was 0.87 in the control test chambers for the third test,
survival and reproduction were both higher in the control test chambers for the third
test; therefore, the pH variation probably did not reduce survival or reproduction.
The pH-adjustment was based on the average measured pH in the lowest tested
concentration of ammonia. The SMCV and the GMCV are <1.45 mg N/L.
Procambarus clarkii (crayfish)
The available data for this species are not used for the reason(s) given in
Appendix 1.
Pteronarcella badia (stonefly)
Thurston et al. (1984a) studied the effect of ammonia on the survival and
emergence of nymphs from two sources for 30 and 24 days. When expressed in
terms of total ammonia nitrogen adjusted to pH=8, the 30-day LC50 for nymphs
from the Gallatin River was about 170 mg N/L, whereas the 24-day LC50 for
nymphs from Rocky Creek was about 70 mg N/L. The degree of development of
the nymphs at the beginning of each test was not determined and there is no
reason to believe that the tested life stage is the one that is most sensitive to
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ammonia. In addition, it is not possible to interpret the data concerning emergence
from either test. The test with nymphs from the Gallatin River might have been
ended before emergence was complete in the control or any other treatment. In the
test with nymphs from Rocky Creek, 25 percent of the nymphs in the control
treatment neither died nor emerged, whereas this percentage was 5 to 15 in the
treatments that contained ammonia. These tests do not allow derivation of a SMCV
for this species, but they imply that this species is resistant to ammonia.
Carasslus auratus (goldfish)
Marchetti (1960) exposed fish for 90 minutes and then observed mortality and
histological effects for up to 42 days, whereas Reichenbach-Klinke (1967) studied
the effects of a one-week exposure on gills and blood. Neither study provided
useful information concerning the SMCV for the goldfish.
Pimephales promelas (fathead minnow)
Thurston et al. (1986) reported similar results from two life-cycle tests that started
with 3 to 5-day-old fry and ended with 60-day-old offspring. The lowest mean
measured DO concentration in any treatment was 6.08 mg/L (72 percent of
saturation) and the lowest calculated fifth percentile of the DO concentrations was
5.16 mg/L (61 percent of saturation). At the highest tested un-ionized ammonia
concentration of 0.93 mg NH3/L, significant mortality occurred throughout the
development of the parental generation. The most sensitive effect was reduction in
egg hatching and the highest concentration that reportedly did not cause a
significant reduction in egg hatching was 0.19 mg NH3/L, but this concentration
caused 33 and 55% reductions in percent hatch. For the purpose of regression
analysis of percent hatch, the tested concentrations and results were so similar in
the two tests that the data were analyzed as replicates of the test concentrations.
In terms of total ammonia nitrogen, the EC20 based on percent hatch was 1.97 mg
N/L at 24.2°C and pH=8.0. However, there are concerns about this test:
1. Effects on survival and weight of F1 fry were uncertain due to high mortality
attributed to handling during cleaning.
2. The eggs were dipped in malachite green daily.
3. Hatchability of the controls was about 50 percent.
4. There was a large difference between the replicate test chambers in the
control-adjusted percent hatch at 0.09 mg NH3/L.
Swigert and Spacie (1983) conducted a 30-day early life-stage test starting with 10
to 18-hour-old embryos. The fifth percentile of the measured DO concentrations
was 6.5 mg/L (79 percent of saturation) and the highest measured DO
concentration was 7.96 mg/L (97 percent of saturation). Both survival and weight
gain were reduced at 30 days and the product of these two (i.e., biomass) was
analyzed using regression analysis. The resulting EC20 was 3.73 mg N/L at
25.1 °C and pH=7.82, which would be 2.92 mg N/L at pH=8.
Mayes et al. (1986) conducted a 28-day early life-stage test in water from the
Tittabawassee River. This water was probably an acceptable dilution water
53
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because it was apparently collected upstream of all known point discharges
(Alexander et al. 1986; James Grant, Michigan Department of Environmental
Quality, personal communication). The lowest and highest measured DO
concentrations were 5.0 and 8.5 mg/L (59 and 101 percent of saturation). Adverse
effects were observed on 28-day survival, but only the highest tested concentration
reduced weight. Regression analysis of the survival data resulted in an EC20 of
5.12 mg N/L at 24.8°C and pH=8.0.
As stated above in the discussion of the effect of temperature on the toxicity of
ammonia, DeGraeve et al. (1987) studied the effect of ammonia on 30-day survival
of juvenile fathead minnows at several temperatures. The tests at 15 and 20°C did
not have concentrations sufficiently high to cause effects, but survival was
significantly decreased at the higher concentrations of ammonia in the tests run at
6, 10, 25, and 30°C. At 30°C, the mean measured DO concentration in most of the
treatments was below 5.5 mg/L, but it was above 60% of saturation in all
treatments. EC20s based on survival were calculated to be 11.9, 13.8, 39, and 39
mg N/L at temperatures of 6.0, 10.0, 25.4, and30.2°C and pHs of 7.83, 7.73, 7.35,
and 7.19, respectively. When adjusted to pH=8, the EC20s are 9.45, 9.72, 19.35,
and 17.54 mg N/L, respectively. Although these EC20s were used to assess the
effect of temperature on the chronic toxicity of ammonia, they are not included in
Table 5 and are not used in the derivation of the SMCV because they indicate that
30-day survival of juveniles is not as sensitive to ammonia as the life-cycle and
early life-stage tests discussed above.
The study of Smith (1984) concerned histopathological examination of lesions on
the test fish and cannot be used to calculate an EC20.
Hermanutz et al. (1987) studied the survival, growth, and reproduction of fathead
minnows in experimental streams. (See Appendix 8, titled "A Field Study Relevant
to the CCC.") Two generations were each exposed for periods of approximately
two months, during which pH averaged 7.5 to 7.7 and temperature averaged
19.6°C. Deleterious effects on biomass were not apparent at or below the highest
tested concentration of ammonia, which was 3.92 mg N/L when adjusted to pH=8.
These results are not included in Table 5 because they are from a field study.
In the 1985 Guidelines (U.S. EPA 1985b), results of early life-stage tests are used
as predictors of results of life-cycle and partial life-cycle tests; comparisons of
these kinds of chronic tests had been reported by McKim (1977) and Macek and
Sleight(1977). Because early life-stage tests are only predictors, results of such
tests are not used when results of life-cycle or partial life-cycle tests are available.
In the present case, however, because of the concerns about the life-cycle test, the
SMCV for the fathead minnow at pH=8 is set equal to 3.09 mg N/L, which is the
geometric mean of the three EC20s from Thurston et al. (1986), Swigert and
Spacie (1983), and Mayes et al. (1986); the range of the three EC20s is only a
factor of 2.6.
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Catostomus commersoni (white sucker)
Reinbold and Pescitelli (1982a) conducted a 31-day early life-stage test starting
with 3-day-old embryos. The concentration of DO averaged 68 to 74 percent of
saturation (6.3 to 6.9 mg/L). No effect on growth or survival was observed at
concentrations of total ammonia nitrogen up to 2.9 mg N/L at pH=8.32 and 18.6°C,
which is equivalent to 4.79 mg N/L at pH=8. As measured by time-to-swimup,
development of larvae was delayed, suggesting that slightly higher concentrations
would have affected growth and/or survival. The results of this test do not provide
sufficient data to allow regression analysis, but the data indicate that the EC20
would be greater than 4.79 mg N/L if an EC20 could be calculated.
Hermanutz et al. (1987) studied survival and growth of juvenile white suckers in
experimental streams. (See Appendix 8.) Two separate tests were started with
individuals whose average weight was 10 g and lasted 88 and 183 days. The
average temperatures in the two tests were 18 and 21 °C. The two highest tested
concentrations caused a slight reduction in biomass. However, juveniles might not
be as sensitive to ammonia toxicity as early life stages. These results are not
included in Table 5 because they are from a field study.
The value of ">4.79 mg N/L" is included in Table 5 and is the GMCV; even though it
is a "greater than" value, it can be used in the calculation of the FCV because it is
not one of the four lowest GMCVs.
Ictalurus punctatus (channel catfish)
Swigert and Spacie (1983) conducted a 30-day exposure starting with newly
hatched larvae that were fewer than 3 hours old. The mean measured DO
concentration was 5.66 mg/L (70 percent of saturation) but the lowest individual
measured concentration was 3.5 mg/L (45 percent of saturation). Reduced growth
was found at total ammonia concentrations of 5.8 mg N/L and above and reduced
survival at concentrations of 21 to 22 mg N/L. In separate tests, they determined
that survival and hatching of embryos were more resistant than survival and growth
of fry. Regression analysis of biomass at the end of the 30-day exposure produced
an EC20 of 11.5 mg N/L at pH=7.76 and 26.9°C. The EC20 adjusted to pH=8 is
8.38 mg N/L. This EC20 is questionable because the lowest measured DO
concentration was below 5.0 mg/L and was below 50 percent of saturation.
Reinbold and Pescitelli (1982a) conducted a 30-day exposure starting with <36-
hour old embryos. The concentration of DO averaged 70 to 76 percent of
saturation (5.7 to 6.2 mg/L). No effect on either percent hatch or fry survival was
found at concentrations up to 11 mg N/L, but reduced growth was found at 5.2 mg
N/L and above, as well as a delay in swimup at concentrations as low as 1 mg N/L.
The EC20 for growth is 12.2 mg N/L at pH=7.80 and 25.8°C. Adjusted to pH=8,
this EC20 is 9.33 mg N/L. However, the percent reduction at the highest tested
concentration was less than 50%, as specified above in the data requirements.
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Colt and Tchobanoglous (1978) and Colt (1978) exposed juveniles for 31 days to
total ammonia nitrogen concentrations ranging from 1.6 to 14.4 mg N/L. The mean
measured DO concentration was 7.6 mg/L (97 percent of saturation) and the
calculated fifth percentile of the DO concentrations was 7.27 mg/L (93 percent of
saturation); the calculated 95th percentile of the DO concentrations was 7.93 mg/L
(101 percent of saturation). Biomass in the control treatment increased tenfold
during the test, but the increases were smaller at ammonia concentrations as low
as 1.6 mg N/L. Because this was a test with juveniles that lasted only 31 days, only
the data concerning mortality will be used. The concentration of 6.81 mg N/L killed
83%, whereas the higher concentration killed 100%. A range is reported for the
concentration of 5.71 mg N/L and so the mean percent mortality is between 28 and
45%. It was reported that the lower concentrations killed 9 of 400 organisms, and
so it is likely that the concentration of 5.02 mg N/L killed no more than 5%.
Therefore, the EC20 at pH=8.35 and 27.9°C is between 5.02 and 5.71 mg N/L;
adjusted to pH=8, the EC20 is between 8.7 and 9.9 mg N/L. Although this EC20 is
included in Table 5, it is not used in the derivation of the SMCV and GMCV
because it is based on survival of juveniles in a 31-day test and therefore is an
upper limit on the SMCV because juveniles might not be as sensitive to ammonia
toxicity as early life stages.
In several tests, each of which consisted of one concentration of ammonia and a
control, Robinette (1976) studied the effect of ammonia on growth of 25 to 30-g
channel catfish for about thirty days at 23 to 26°C. No information was reported
concerning survival of the test fish. A concentration of total ammonia nitrogen of
2.7 mg N/L at pH=7.6 caused fish to gain weight faster than the control fish. In
contrast, concentrations of 3.5 and 3.6 mg N/L at pH=7.8 caused fish to lose weight
while the controls were gaining weight. Adjusted to pH=8, these concentrations
would be 1.7, 2.7, and 2.8 mg N/L, respectively. Because these tests studied
growth of juveniles for only 30 days, the results are not included in Table 5.
Bader (1990) and Bader and Grizzle (1992) reported that ammonia reduced
growth, but the concentration of ammonia in the controls was substantial.
DeGraeve et al. (1987) studied the effect of ammonia on survival and growth of
juveniles for thirty days. Some of the test organisms were treated with acriflavine
up to two days prior to the beginning of the test. In addition, the mean measured
DO concentration was below 5.5 mg/L and below 60 percent of saturation in some
of the treatments. Mitchell and Cech (1983) reported that ammonia did not damage
gills unless residual chlorine was present. Soderberg et al. (1984) studied the
culture of channel catfish in ponds and found that the ambient concentration of
ammonia caused gill lesions, but did not affect survival or growth. Results of these
tests are not included in Table 5.
Hermanutz et al. (1987) studied survival and growth of juvenile channel catfish in
experimental streams. (See Appendix 8.) Three separate tests lasted from 36 to
177 days and were started with individuals whose average weights ranged from 6
to 19 g. Average temperatures in the three tests were 17 to 21 °C. Both of the
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longer tests showed monotonic, substantial reductions in biomass; these results
are in reasonable agreement with the results of the laboratory tests. However,
juveniles might not be as sensitive to ammonia toxicity as early life stages are.
These results are not included in Table 5 because they are from a field study.
Although there are problems with the early life-stage tests by Swigert and Spacie
(1983) and Reinbold and Pescitelli (1982a), the EC20s are similar. Therefore, the
channel catfish SMCV at pH=8 is 8.84 mg N/L, which is the geometric mean of the
two EC20s. The data of Colt and Tchobanoglous (1978) and Robinette (1976)
support a SMCV of this magnitude. The GMCV is also 8.84 mg N/L.
Oncorhynchus clarki (cutthroat trout)
Thurston et al. (1978) obtained 29-day LCSOs of 16.4 and 15.9 mg N/L with fish
whose average weights were 3.3 and 3.4 g, respectively; the 96-hr LCSOs were 1.2
and 1.7 times higher than the 29-day LCSOs. In two other tests they obtained 36-
day LCSOs of 23.7 and 24.4 mg N/L with fish whose average weight was 1.0 g; no
fish died after day 29. The tests were conducted at 12.2 to 13.1 °C and all four of
the LCSOs are expressed as total ammonia nitrogen at pH=8.0. The mean
measured DO concentrations for the various tests ranged from 8.2 to 8.6 mg/L (77
to 82 percent of saturation). The lowest and highest measured DO concentrations
were 7.4 and 9.2 mg/L (70 and 87 percent of saturation). EC20s cannot be
calculated, but would be lower than the geometric mean of 19.7 mg N/L. The
SMCV might be substantially lower than 19.7 mg N/L because this test was not
conducted with an early life stage. In all four of the tests, there was a negative
correlation between the concentration of ammonia and weight gain, but this might
have been a temporary effect. Histological examinations were performed at the
end of the tests. The EC20 of <19.7 mg N/L is included in Table 5, but this value
cannot be used in the calculation of a SMCV.
Oncorhynchus gorbuscha (pink salmon)
Rice and Bailey (1980) exposed embryos and alevins of pink salmon for 61 days to
concentrations of total ammonia nitrogen ranging from 0.07 to 13.6 mg/L at pH=6.4
and 4°C. The only chronic test began sometime after hatch and ended when the
alevins emerged (i.e., at the beginning of swimup); therefore the test did not
include effects of ammonia on the growth and survival of fry after feeding started.
In addition, no information was given concerning survival to the end of the test in
the control or any other treatment. At the higher tested concentrations, the weight
of emerging alevins was significantly reduced, relative to the controls, by as much
as 22% at 11.2 mg/L. This would be equivalent to about 4.1 mg N/L at pH=8. Size
at emergence was said to be important because smaller fry are less capable of
surviving in the environment because they have less swimming endurance and are
selectively preyed upon by larger predators. This test did not provide data
concerning survival and is not an early life-stage test because it began after hatch;
therefore, this test did not provide a useful EC20 and is not included in Table 5.
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Oncorhynchus kisutch (coho salmon)
Buckley et al. (1979) exposed fish whose average wet weight was 3.4 g for 91 days
to study effects of ammonia on blood. The highest tested concentration of 47 mg
N/L killed only three percent of the fish. The EC20 is >47 mg N/L, but this not
useful information about the SMCV because there is no reason to believe that the
tested life stage is the one that is most sensitive to ammonia. This test is not
included in Table 5 because it does not provide useful information concerning the
SMCV for this species.
Oncorhynchus mykiss (Salmo gairdneri) (rainbow trout)
Many investigators have reported results of chronic tests conducted on ammonia
with rainbow trout, but the most ambitious chronic test was the five-year test
conducted by Thurston et al. (1984b). In this test the initial fish were exposed
through growth, maturation and reproduction, the next generation through hatch,
growth, maturation, and reproduction, and the third generation through hatch and
survival of the young. The mean measured DO concentration was 7.43 mg/L (65
percent of saturation) and the lowest calculated fifth percentile of the measured DO
concentrations in the various treatments was 5.9 mg/L (51 percent of saturation).
Measured temperatures ranged from 7.5 to 10.5°C and the tested concentrations of
total ammonia nitrogen ranged from 1.1 to 8.0 mg N/L at pH=7.7. When adjusted
to pH=8, the range is 0.77 to 5.4 mg N/L. All of the fish used to start the test came
from one pair of adults of the Ennis strain. In addition, the important data for each
life stage are so variable that it is not possible to discern whether there is a
concentration-effect curve. Despite the variability, it can be inferred that the EC20
cannot be much lower than the highest tested concentration because severe
effects were not apparent at any tested concentration; if the EC20 was much lower
than the highest tested concentration, this concentration would have caused severe
effects.
Also using fish from the Ennis strain, Burkhalter (1975) and Burkhalter and Kaya
(1977) reported a 21 -day LC50 of 39.6 mg N/L for embryos and sac fry and
interpolation off a graph indicates a 42-day LC50 of 33.6 mg N/L, based on total
ammonia nitrogen, at 9.5 to 12.5°C and pH=7.5, assuming either no control
mortality or adjustment for control mortality. When adjusted to pH=8, the LCSOs
would be 22.0 and 18.7 mg N/L, respectively, but LC20s would be lower than
LCSOs. The measured DO concentrations were all above 8 mg/L (72 percent of
saturation). The test began within 24 hours of fertilization, continued to the
beginning of feeding, and found retardation of development and growth of very
young fish, similar to the tests discussed above with the pink salmon (Rice and
Bailey 1980). Thurston et al. (1984b) speculated that they did not observe the
reduced growth reported by Burkhalter and Kaya (1977) because of compensation
during the next several months of the longer exposure. Indeed, Burkhalter and
Kaya (1977) reported compensation at the lowest tested concentration.
Contrasting information concerning EC20s is provided by the early life-stage tests
conducted by Solbe and Shurben (1989) and Calamari et al. (1977,1981). Both
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tests began within 24 hours after fertilization and lasted for 72 to 73 days until the
fry had been feeding for about 30 days.
1. Solbe and Shurben (1989) reported that the dry weight of the test organisms
varied little between treatments. The test was conducted at pH=7.52 and an
average temperature of 14.9°C. The DO concentration equaled or exceeded
76 to 95 percent of saturation during various portions of the test. The four
highest concentrations of ammonia killed 78 to 99 percent. The fifth and lowest
tested concentration of total ammonia nitrogen was 2.55 mg N/L and it reduced
survival by 67 percent; this would correspond to 1.44 mg N/L at pH=8, and the
LC20 would be lower. These authors demonstrated that exposure to ammonia
should begin soon after fertilization. When exposure began within 24 hours
after fertilization, 26 mg N/L killed 98 percent of the embryos, whereas when
exposure began 24 days after fertilization, 26 mg N/L killed only 3 percent of
the embryos and killed only 40 percent in a 49-day exposure.
2. Calamari et al. (1977,1981) conducted an early life-stage test, but did not
report any information concerning weight, although, as stated above, Solbe and
Shurben (1989) reported no effect on weight during their early life-stage test.
The DO concentration was over 80 percent of saturation. For total ammonia
nitrogen at pH=7.4, Calamari et al. (1977,1981) obtained a 72-day LC50 of 8.2
mg N/L at 14.5°C. They also reported that adjusted mortalities were 15 and 23
percent at 1.5 and 3.7 mg N/L, respectively, and that higher tested
concentrations killed more than 50 percent of the test organisms. Because
Calamari et al. did not report the actual percentage killed at the higher tested
concentrations, regression analysis could not be applied; semilog interpolation
between 1.5 and 3.7 mg N/L produced an LC20 of 2.6 mg N/L, which would
correspond to 1.34 mg N/L at pH=8.
Both Calamari et al. (1977,1981) and Solbe and Shurben (1989) found that longer
exposures of embryos and fry resulted in much lower LCSOs than 96-hour
exposures.
Several investigators reported results concerning the effect of total ammonia
nitrogen on long-term survival:
1. Thurston and Russo (1983) reported five 35-day LCSOs that were determined
using fish whose average initial weights were 0.7 to 10 g. The 35-day LCSOs
were 27.9 and 36.1 mg N/L for fish whose average weights were 3.7 and 9.7 g,
respectively. The 35-day LCSOs were 32.4, 34.5, and 37.0 mg N/L for fish
whose average weights were 0.7 to 3.3 g; when adjusted to pH=8, the
geometric mean of these three 35-day LCSOs with the smaller fish was 26.4 mg
N/L.
2. Broderius and Smith (1979) reported that 16.2 mg N/L killed 30 percent of fry in
30 days at 10°C and pH=7.95, which corresponds to 15.1 mg N/L at pH=8.
3. Daoust and Ferguson (1984) reported that 23.3 mg N/L did not kill any
fingerlings in 90 days at pH=7.93, which would correspond to 21.1 mg N/L at
pH=8. However, some of the fish that exhibited clinical signs during the
exposure were removed for examination during the test. The swimming and
feeding of some fish were affected for a while, but the fish recovered.
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This variety of results might be due to differences in the size or age of the test
organisms.
Several other chronic tests did not provide information that could be used in the
derivation of a SMCV. Fromm (1970), Reichenback-Klinke (1967), and Smart
(1976) exposed fish to study the effects of ammonia on gills and blood. In a test
reported by Smith and Piper (1975), exposed fish had abnormal tissues, but fish
placed in clean water for 45 days at the end of the test had normal tissues. When
Soderberg et al. (1983) studied the culture of rainbow trout in ponds, parasitic
epizootics caused mortalities. The Ministry of Technology (1967) reported the
effect of ammonia on percent survival in a 90-day test, but did not report the age or
size of the fish or the temperature or the pH of the water. Samylin (1969)
conducted tests in water from the Vyg River, with some of the exposures being
conducted in Petri dishes. Schulze-Wiehenbrauck (1976) found that growth of
juveniles at 10°C and pH=8 was reduced during two-week exposures to a total
ammonia nitrogen concentration of 2.26 mg N/L, but the decrease was completely
compensated for during the next three or four weeks. Smith (1972) reported that as
long as the DO concentration was maintained at 5 mg/L or greater, growth of
rainbow trout was not significantly reduced until average total ammonia
concentrations reached 1.6 mg/L.
Hermanutz et al. (1987) studied survival and growth of juvenile rainbow trout in
experimental streams. (See Appendix 8.) Three separate tests were conducted
with individuals whose average initial weights were 7 to 11 g. The tests lasted from
28 to 237 days, with the 237-day test including an entire winter. Average
temperatures in the three tests ranged from 5.9 to 10.6°C, whereas pH averaged
7.7 to 8.4. Reductions in biomass were consistently observed at concentrations
greater than or equal to 2.29 mg N/L when adjusted to pH=8. However, juveniles
might not be as sensitive to ammonia toxicity as early life stages. These results
are not included in Table 5 because they are from a field study.
The early life-stage test by Calamari et al. (1977,1981) produced a total ammonia
nitrogen LC20 of 1.34 mg N/L at pH=8, whereas Solbe and Shurben (1989)
indicate that the LC20 might be lower. In contrast, both Thurston et al. (1984a) and
Burkhalter and Kaya (1977) found no indication of severe mortality of young fish at
higher concentrations. Exposure was continuous for several generations in the test
of Thurston et al. (1984b), whereas exposure began within 24 hours of fertilization
in the other three tests. Because of the concerns about some of the tests, the
differences among the results, and the fact that some of the results are either
"greater than" or "less than" values, even though the various results are included in
Table 5, a SMCV is not derived for rainbow trout; instead, the results of the chronic
tests will be used to assess the appropriateness of the CCC.
Oncorhynchus nerka (sockeye salmon)
Rankin (1979) exposed embryos of sockeye salmon for 62 days from fertilization to
hatch; the tested concentrations of total ammonia nitrogen ranged from 2.13 to 87
60
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mg N/L at 10°C. The DO concentration was reported to be at saturation. This test
ended as soon as the embryos hatched, and so hatchability was the only
toxicological variable studied. The percentage of the embryos that hatched was
63.3% in the controls, but was 49% at the lowest tested concentration (2.13 mg
N/L) and was 0% at 8.1 mg N/L and above. The concentration of 2.13 mg N/L at
pH=8.42 corresponds to 4.16 mg N/L at pH=8. Thus the EC20 at pH=8 is less than
4.16 mg N/L. Because the effects on newly hatched fish were not studied, the
SMCVis<4.16mgN/L
Oncorhynchus tshawytscha (chinook salmon)
Burrows (1964) exposed fingerlings for six weeks at 6 and 14°C to three
concentrations of ammonia and a control treatment to study effects on gills at
pH=7.8. There was no recovery in three weeks in clean water at 6°C, but there
was recovery at 14°C. At both temperatures, no significant mortality occurred
during exposure to the highest tested concentration of 0.57 mg N/L or for three
weeks afterward in clean water. No information is given concerning the DO
concentration during the exposures, and there is no reason to believe that the
tested life stage is the one that is most sensitive to ammonia.
Tests conducted by Sousa et al. (1974) suggest that chinook salmon tolerate
higher concentrations of ammonia when pH is decreased and salinity is increased.
However, there was no control treatment, no information was given concerning the
DO concentration, temperature was not controlled, and the fish were given an
antibiotic.
These tests are not included in Table 5 because they do not provide useful
information concerning the SMCV for this species.
A GMCV is not derived for Oncorhynchus because the available data do not
provide an adequate basis for a useful conclusion concerning the GMCV.
Salmo trutta (brown trout)
Carline et al. (1987) exposed brown trout for twelve months to dilutions of effluent
from a sewage treatment plant. Survival, growth, swimming performance, and
degree of damage to gills were studied, but no information was obtained
concerning effects on embryos, newly hatched fish, or reproduction. No data from
this test are included in Table 5 because this test does not provide useful
information concerning the SMCV for this species.
Lepomis cyanellus (green sunfish)
Reinbold and Pescitelli (1982a) conducted a 31-day early life-stage test that
started with <24-hour-old embryos. No information was reported concerning the
DO concentration but it averaged 70 to 76 percent of saturation (5.7 to 6.2 mg/L) in
a similar test in the same report with another fish species at about the same
temperature. The weight data were not used in the calculation of an EC20 because
the fish were heavier in chambers containing fewer fish, which indicated that weight
61
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was density-dependent. Although overflows resulted in loss offish from some
chambers, survival was 96 percent in one of the chambers affected by overflow,
indicating that the survival data were either adjusted or not affected by the
overflows. Survival to the end of the test was reduced at total ammonia nitrogen
concentrations of 6.3 mg N/L and above and regression analysis of the survival
data calculated an EC20 of 5.84 mg N/L at pH=8.16 and 25.4°C. Adjusted to
pH=8, the EC20 is 7.44 mg N/L.
McCormick et al. (1984) conducted a 44-day early life-stage test, starting with <24-
hour-old embryos. The mean measured DO concentration was 7.9 mg/L (91
percent of saturation) and the calculated fifth percentile of the measured DO
concentrations was 7.7 mg/L (88 percent of saturation). No effect was found on
percent hatch, but reduced survival and growth occurred at concentrations of 14
mg N/L and above. Although survival in one control test chamber and in the low
concentrations of ammonia averaged about 40 percent and was only 10 percent in
the other control chamber, the concentration-effect curve was well defined.
Regression analysis of biomass calculated an EC20 of 5.61 mg N/L at pH=7.9 and
22.0°C. This EC20 was obtained with the 10 percent used in the regression
analysis. An EC20 of 5.51 mg N/L was obtained if the 10 percent was not used; the
two EC20s are similar partly because the weight given to each treatment was
inversely related to the variance for the treatment, which meant that the control
treatment was given a low weight in the regression analysis. Adjusted to pH=8, the
EC20 calculated using all of the data is 4.88 mg N/L.
Jude (1973) found that growth of juveniles weighing 4 to 16 g each for 40 days was
proportional to temperature at 13, 22, and 28°C. In a second test, the effect of
ammonia on survival and growth of 10 to 14-g juveniles was studied for 20 days.
Too few fish died to allow calculation of an EC20. Neither of these tests provided
results that can be included in Table 5.
Adjusted to pH=8, the EC20 of 7.44 mg N/L from Reinbold and Pescitelli (1982a)
agrees quite well with the EC20 of 4.88 mg N/L from McCormick et al. (1984). It is
possible that the second value is lower because it was based on survival and
growth, whereas the first value was based only on survival. Even though there
were experimental problems with both tests, the results of the tests agree well and
therefore the geometric mean (6.03 mg N/L) of the two EC20s is used as the
SMCV.
Lepomis macrochirus (bluegill)
Smith et al. (1984) conducted a 30-day early life-stage test, starting with <28-hour-
old embryos. No information was reported concerning the DO concentration, but
the flow-rate was high. The values reported in their Table 1 as standard deviations
on the pH appear excessively large; it is likely that they were not calculated
correctly, because, as explained in footnote d, the mean pH was calculated by
conversion of pH to H+ (i.e., hydrogen ion) concentration. Other tests conducted
on ammonia in the same laboratory at about the same time reported much less
62
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variation in pH. For example, McCormick et al. (1984) reported that the 95%
confidence interval on the experiment-wide pH was 7.8 to 8.0. Broderius et al.
(1985) calculated average pH by converting to hydrogen ion concentration, but
reported small standard deviations and ranges for four acute tests and four chronic
tests.
Smith et al. (1984) found no significant reduction in percent hatch up to a total
ammonia nitrogen concentration of 37 mg N/L, but hatched larvae were deformed
at this concentration and died within six days. At the end of the test, survival and
growth at 1.64 mg N/L were near values for the controls, but were greatly reduced
at 3.75 to 18 mg N/L. Regression analysis of biomass calculated an EC20 of 1.85
mg N/L at pH=7.76 and 22.5°C. The EC20 adjusted to pH=8 is 1.35 mg N/L.
Diamond et al. (1993) conducted two chronic tests. The test at 12°C is discussed
in Appendix 1. The data sheets for the test at 20°C indicate that this test studied
the effect of ammonia on survival and growth of bluegills for 21 days. (The
durations of the chronic tests with the bluegill at 12 and 20°C are switched in Table
1 in the publication.) The test at 20°C was started with bluegills that were less than
98-days old, were less than 1 inch (2.5 cm), and averaged 0.11 to 0.15 g. The
highest tested concentration of total ammonia nitrogen was 64 mg N/L, which
caused 30% mortality at the test pH of 7.3; most of the deaths occurred in the last
two days of the test. Adjusted to pH=8, the highest tested concentration was 31 mg
N/L as total ammonia nitrogen, which is in the range of the adjusted 96-hr LCSOs
reported in Table 1 of the 1984/1985 ammonia criteria document. This test is not
very useful because it lasted for only 21 days and mortality began occurring near
the end of the test. Neither of these tests provides results that can be included in
Table 5.
Hermanutz et al. (1987) studied survival and growth of the juvenile bluegills in
experimental streams. (See Appendix 8.) The individual weights averaged 2.2 g at
the beginning and the test duration was 90 days. The mean pH and temperature
were 8.2 and 21.1 °C, respectively. A substantial effect on biomass was apparent
only at the highest concentration, which was 9.5 mg N/L when adjusted to pH=8.
These juvenile bluegills were not particularly sensitive compared to older life
stages of other species tested during this study. However, juveniles apparently are
not as sensitive to ammonia toxicity as the early life stages tested by Smith et al.
(1984). These results are not included in Table 5 because they are from a field
study.
The SMCV for the bluegill is 1.35 mg N/L, and the GMCV of 2.85 mg N/L for
Lepomis is calculated as the geometric mean of the two SMCVs (6.03 and 1.35 mg
N/L).
Micropterus dolomieu (smallmouth bass)
As stated above in the discussion of the effect of pH on the toxicity of ammonia,
Broderius et al. (1985) conducted 32-day early life-stage tests at four pHs at
63
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22.3°C, starting with embryos near hatch. The mean measured DO concentration
was 7.72 mg/L (89 percent of saturation); the lowest and highest measured DO
concentrations were 7.1 and 8.3 mg/L (81 and 96 percent of saturation). Survival
of embryos and fry within the first week was not affected by ammonia, except at the
highest concentration at the highest pH, although effects on these life stages might
have been reduced due to the exposure not starting until just prior to hatch. In all
tests, growth and survival of older fry were reduced at higher concentrations and
regressions of biomass resulted in EC20s of 9.61, 8.62, 8.18, and 1.54 mg N/L at
pHs of 6.60, 7.25, 7.83, and 8.68, respectively. Adjusted to pH=8, these EC20s are
3.57, 4.01, 6.50, and 4.65 mg N/L, with a geometric mean of 4.56 mg N/L, which is
the SMCV and the GMCV.
Stizostedion vitreum (walleye)
Reinbold and Pescitelli (1982a) could not conduct a successful early life-stage test
because only 20% of the newly hatched fish survived.
Hermanutz et al. (1987) studied survival and growth of juvenile walleyes in
experimental streams. (See Appendix 8.) A 46-day test was conducted at an
average temperature of 24°C and was started with yearlings averaging 100 g initial
weight. A second test at an average temperature of 17°C was started with young-
of-year averaging 19 g initial weight and lasted 43 days. Adjusted to pH=8,
concentrations of 2.0 to 3.7 mg N/L somewhat reduced walleye biomass, whereas
concentrations of 9.5 to 13.3 mg N/L completely eliminated walleye from the
streams. However, juveniles might not be as sensitive to ammonia toxicity as early
life stages. These results are not included in Table 5 because they are from a field
study.
Rana pipiens (leopard frog)
The available data for this species are not used for the reason(s) given in
Appendix 1.
Hyla crucifer (spring peeper)
The available data for this species are not used for the reason(s) given in
Appendix 1.
64
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Table 5. EC20s from Acceptable Chronic Tests8
Species
Musculium
transversum
Ceriodaphnia
acanthina
Ceriodaphnia
dubia
Daphnia magna
Hyalella
azteca
Pimephales
promelas
Reference
Anderson et al.
1978
Sparks and
Sandusky 1981
Mount 1982
Willingham
1987
Nimmo et al.
1989
Gersich et
al. 1985
Reinbold and Pescitelli
1982a
Borgmann 1994
Thurston et al.
1986
Swigert and
Spacie 1983
Mayes et al.
1986
Test and
Effect"
42-d Juv
Survival
42-d Juv
Survival
LC
Reproduction
7-dLC
Reproduction
7-d LC
Reproduction
21 -d LC
Reproduction
21 -d LC
Reproduction
10-wkLC
Reproduction
LC
Hatchability
30-d ELS
Biomass
28-d ELS
Survival
Temp.
(C)
23.5
21.8
24.5
26.0
25.
19.8
20.1
25.
24.2
25.1
24.8
PH
8.15
7.80
7.15
8.57
7.8
8.45
7.92
7.94
8.0
7.82
8.0
EC20C at
test pH
& Temp.
(mg N/L)
5.82
1.23
44.9
5.80
15.2
7.37
21.7
<1.58
(EC50)
1.97
3.73
5.12
EC20C
at pH=8
&25°C
(mg N/L)
6.63
0.77
19.1
15.6
11.6
10.8
14.1
<1.45
1.97
2.92
5.12
SMCVC
at pH=8
&25°C
(mg N/L)
.2b
19.1
13.5
12.3
<1.45
3.09
GMCVC
at pH=8
&25°C
(mg N/L)
<2..2b
16.1
12.3
<1.45
3.09
65
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Species
Catostomus
commersoni
Ictalurus
punctatus
Oncorhynchus
clarki
Oncorhynchus
mykiss
Oncorhynchus
nerka
Reference
Reinbold and Pescitelli
1982a
Swigert and Spacie
1983
Reinbold and Pescitelli
1982a
Colt and
Tchobanoglous 1978
Thurston et al.
1978
Thurston etal. 1984b
Burkhalter and
Kaya1977
Solbe and
Shurben 1989
Calamari et
al. 1977,1981
Rankin 1979
Test and Effect13
30-d ELS
Biomass
30-d ELS
Biomass
30-d ELS
Weight
30-d Juv
Survival
29-d Juv
Survival
5-year LC
42-d ELS
Survival
73-d ELS
Survival
72-d ELS
Survival
62-d Embryos
Hatchability
Temp.
(C)
18.6
26.9
25.8
27.9
12.2-
13.1
7.5-
10.5
9.5-
12.5
14.9
14.5
10.
PH
8.32
7.76
7.80
8.35
8.0
7.7
7.5
7.52
7.4
8.42
EC20C at
test pH
& Temp.
(mg N/L)
>2.9
11.5
12.2
<5.02-
<5.71
<19.7
>«8.0
<33.6
<2.55
2.6
<2.13
EC20C
at pH=8
&25°C
(mg N/L)
>4.79
8.38
9.33
<8.7-
<9.9d
<19.7d
>«5.4d
<18.7d
<1 .44d
1.34d
<4.16
SMCVC
at pH=8
&25°C
(mg N/L)
>4.79
8.84
—
<4.16e
GMCVC
at pH=8
&25°C
(mg N/L)
>4.79
8.84
—
66
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Species
Lepomis
cyanellus
Lepomis
macrochirus
Micropterus
dolomieu
Reference
Reinbold and Pescitelli
1982a
McCormick et al. 1984
Smith et al.
1984
Broderius et
al. 1985
Broderius et
al. 1985
Broderius et
al. 1985
Broderius et
al. 1985
Test and Effect13
30-d ELS
Survival
30-d ELS
Biomass
30-d ELS
Biomass
32-d ELS
Biomass
32-d ELS
Biomass
32-d ELS
Biomass
32-d ELS
Biomass
Temp.
(C)
25.4
22.0
22.5
22.3
22.3
22.3
22.3
PH
8.16
7.9
7.76
6.60
7.25
7.83
8.68
EC20C at
test pH
& Temp.
(mg N/L)
5.84
5.61
1.85
9.61
8.62
8.18
1.54
EC20C
at pH=8
&25°C
(mg N/L)
7.44
4.88
1.35
3.57
4.01
6.50
4.65
SMCVC
at pH=8
&25°C
(mg N/L)
.(Jo
1.35
.OO
GMCVC
at pH=8
&25°C
(mg N/L)
2.85
.OO
a An EC20 is assumed for a stonefly but is not given in this table (see text concerning calculation of the
CCC).
b Juv = juvenile; LC = life cycle; ELS = early life stage.
c Total ammonia nitrogen.
d Not used in the derivation of a SMCV (see text).
e Not used in the derivation of a GMCV (see text).
67
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Seasonality of Chronic Toxicity Endpoints
Dischargers that use biological treatment of ammonia are likely to find it most difficult to
meet water quality-based effluent limits for ammonia when the temperature is low. This
has raised questions about whether criteria based on toxicity tests conducted mainly at
warm temperatures appropriately define concentrations that are needed for aquatic life
protection under cold-weather conditions. The effect of temperature on chronic EC20s
was discussed in an earlier section of this document. This section will consider the
relevance of particular chronic endpoints at various times of the year.
The CMC is appropriate during all portions of the year because the organisms (i.e.,
juvenile and adult fish) and effects (i.e., survival) on which it is based are relevant
during all portions of the year and because available data indicate that these endpoints
are largely independent of temperature. The CCC, however, is based in part on
endpoints that might not be of concern during cold-season conditions (fish early life
stages, Hyalella reproduction). Consequently, it is also necessary to consider the
effect of seasonality on the chronic endpoint selection.
Endpoint Selection for Fish
Two of the four most sensitive GMCVs used to quantify the chronic criterion are for fish
genera - Pimephales and Lepomis. The next most sensitive genera were also fish
(Micropterus, Ictalurus, Catostomus), which could become more important in the criteria
derivation if one or more of the four most sensitive genera tested at 20-25°C are
sufficiently less sensitive at low temperatures. All of these GMCVs were based on
early-life stage or life-cycle tests. Because the sensitive test endpoints involved
survival and/or growth of very young fish that are present only at temperatures at which
reproduction occurs, these GMCVs are not necessarily appropriate for lower
temperatures. Seasonal criteria should consider the likely response of life stages
present during particular seasons.
The appropriate tests for deriving criteria when early life stages are not present would
be long chronic tests of survival and growth with older life stages, conducted at a wide
range of temperatures. Such data are not available. Life cycle tests with fathead
minnows conducted by Thurston et al. 1986 provide survival and growth data for the
parental generation from a few days of age through spawning (150-200 days), but only
at one temperature (24°C). DeGraeve et al. (1987) conducted chronic exposures with
fathead minnows at four different temperatures, but the duration was only 30 days.
Other chronic exposures with juvenile warm-water fish are limited by both being only at
high temperature and having relatively short durations.
Thus, available chronic data for juvenile and adult fish are somewhat limited with
respect both to the duration of the tests and the temperature range of the tests. With
regard to the limited temperature range, the importance and likely magnitude of
68
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temperature effects was discussed above. Despite various uncertainties, available
information can serve as the basis for reasonable assumptions on how tests at warm
temperatures can be used to derive criteria applicable to lower temperatures. For the
calculations in this document, it will be assumed that for fish the EC20 for a particular
type of effect does not vary with temperature.
With regard to the duration issue, it is not justified to simply assume that 30-day
exposures are long enough to assess chronic effect concentrations for juvenile and
adult fish, even though this is the duration accepted for early life stage tests for warm
water fish. For such early life stage tests, a test duration of 30 days is considered
adequately long because it encompasses a wide range of developmental stages that
likely include the most sensitive endpoints, and because empirical comparisons have
suggested these tests are often nearly as sensitive as longer tests. However, this is
not necessarily true if the tests of interest address less sensitive endpoints involving
juvenile and adult survival and growth. For example, in the fingernail clam tests the
effect concentrations (juvenile survival) at 6 weeks were significantly lower than those
at 4 weeks. Also, for 90-day ammonia exposures to rainbow trout reported by Ministry
of Technology (1967), most of the observed mortality occurred after 30 days.
With regard to the warm water fish of concern here, the effect of duration is most
evident in the fathead minnow life cycle test by Thurston et al. (1986). Starting with
larvae a few days old, these investigators reported survival after 30 days, after 60 days,
and just prior to spawning (155 days in one test and 200 days in another). The first 30
days falls under the realm of early life stage tests and will be ignored here. The
mortality between 30 and 60 days covers a fish age and exposure duration similar to
short chronic tests with juveniles, such as that of DeGraeve et al. (1987). During that
time, mortality was approximately 25% at a total ammonia concentration of 18 mg N/L
and averaged <2% at lower concentrations. This suggests an LC25 of about 18 mg
N/L, which is consistent with the results of DeGraeve et al. at temperatures of 25 and
30°C. But mortality continued after this 30 day period. By spawning time, total
mortality at 18 mg N/L reached about 67% and a mortality of about 10% was observed
at 9 mg N/L, in contrast to an average mortality of <2% at lower concentrations. The
estimated LC25 for this longer exposure was 12.3 mg N/L. This is a factor of 1.5 less
than the LC25 at the shorter exposure. This factor will be used to adjust 30-day effect
concentrations from other tests with juvenile fish to be more reflective of long chronic
exposures.
With the issues of test temperature and test duration so addressed, chronic, "non-
early-life-stage" effect concentrations for different species of warm-water fish can be
assessed as follows:
Fathead minnow
DeGraeve et al. (1987) exposed juvenile fathead minnows to ammonia for 30 days
at various temperatures, as previously discussed. When corrected to pH=8 using
the chronic pH relationship and divided by the factor of 1.5 to adjust to longer
exposures, chronic LC20s from this study at pH=8 are estimated to be 6.5, 8.5,
69
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12.9, and 10.6 mg N/L at 6, 10, 25, and 30°C, respectively. If temperature effects
are assumed to be negligible, the geometric average of these four values, 9.3 mg
N/L, can be used as a representative value for long-term LC20s for juveniles of this
species. If there were concern about better representing low temperature
conditions, the average of the two values at lower temperatures, 7.5 mg N/L, could
be used.
As discussed above, Thurston et al. (1986) conducted two life-cycle exposures of
fathead minnows to ammonia at 24°C and pH«8, starting with 3- to 5-day-old
larvae. For the exposure of fish from 30 day old through spawning, the data and
regression indicate a long-term chronic LC20 of 11.4 mg N/L, similar to the higher
temperature results from DeGraeve et al. (1987).
While the effects concentrations summarized here can be directly used to set a
GMCV for fathead minnows in the absence of early life stages, they can also be
used to develop ratios to other endpoints. Such ratios might be useful for
estimating GMCVs for other species for which chronic tests with juveniles and
adults have not been conducted. One useful comparison would be to the early life
stage and life cycle test results. For fathead minnows, Table 5 listed the GMCV at
pH=8 as 3.09 mg N/L based on the results of two early life stage tests and one life
cycle test. Using the estimate of 9.3 mg N/L from above as representative of what
the GMCV should be when early life stages are absent, this represents a factor of 3
increase in effect concentration. A second comparison that is useful is to the acute
toxicity of this species. The GMAV for fathead minnow at pH=8 is listed as 43.6 mg
N/L in Table 4. This is 4.7-fold higher than the 9.3 mg N/L derived here for chronic
toxicity of older life stages from the DeGraeve et al. (1987) data.
Channel catfish
One of the chronic tests with channel catfish was a 31-day test of juvenile growth
and survival by Colt and Tchobanoglous (1976). These authors did not provide
complete data on survival at each concentration, but the information provided
indicated that (a) at 5.0 mg N/L (total ammonia) and below, mortality averaged
about 2%, (b) at 5.7 mg N/L mortality was between 28 and 45% (based on a
reported range of 11 to 62% mortality in three replicates), (c) at 6.8 mg N/L
mortality was 83%, and (d) at 9.5 mg N/L and above mortality was 100%. This
indicates that the LC20 was between 5.0 and 5.7 mg N/L (at pH=8.35 and T=28°C),
or between 8.7 and 9.9 mg N/L when adjusted to pH=8 based on the chronic pH
relationship. While this test was not used in deriving the GMCV, it did indicate that
juvenile channel catfish were as sensitive as early life stages and was included as
support that the GMCV derived from the early life stage tests was reasonable. In
fact, if the factor of 1.5 is applied to this number to make it more applicable to
longer chronic exposures, the GMCV would be around 6.2 mg N/L - below the
GMCV (8.84 mg N/L) derived from the early life stage tests. Because the 1.5
adjustment factor has questionable applicability to channel catfish, the GMCV from
Table 5 is the most appropriate number to use for both early-life-stage and juvenile
sensitivity.
70
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Juvenile channel catfish show effects on growth at even lower concentrations. The
EC20 for growth from the study of Colt and Tchobanoglous was 2.39 mg N/L at
pH=8.35, or 4.2 mg N/L when adjusted to pH=8. Robinette (1976) also found
growth effects at similar concentrations. In separate 27- and 29-day exposures
with 3.5-3.6 mg N/L total ammonia (pH=7.8), he found fingerling channel catfish to
lose weight, while control fish gained weight. These results were not used in the
criterion derivation, because growth effects on juvenile fish are sometimes
transient, so that short chronic tests with such fish might overestimate risk.
Nonetheless, this does indicate effects of ammonia at rather low concentrations.
The relationship of these effects on juvenile channel catfish to effects observed in
early life stages are quite different than in fathead minnow. Whereas fathead
minnow juvenile tests show about three-fold less sensitivity than early life stage
tests, channel catfish juvenile tests seem to be as, or more, sensitive than the
available early life stage tests. This might be in part due to comparisons across
different studies, but the differences are so great that it does suggest some
fundamental interspecies differences in this regard. It also makes uncertain the
development of a ratio which can be applied to early life stage tests of other
species to estimate what GMCVs should be for later life stages. In contrast, the
relationship of the juvenile chronic test results to acutely lethal concentrations
seem more consistent. In Table 4, the channel catfish SMAV at pH=8 is listed as
34.4 mg N/L total ammonia. This is a factor of 5.5 greater than the long-term
juvenile LC20 estimated above and a factor of 3.9 greater than the GMCV for
catfish in Table 5, both similar to the factor of 4.7 for fathead minnow.
Bluegill sunfish
Bluegills are the only other nonsalmonid species for which juveniles have been
chronically exposed in the laboratory. Diamond et al. (1993) conducted a 21-day
exposure with young bluegill (ca. 150 mg) and reported significant mortality (30%)
and reduced growth at 64 mg N/L total ammonia (pH=7.4, T=22°C). A comparison
of this value to effect concentrations discussed above requires adjustment for (a)
the low pH, (b) the short duration, and (c) mortality>20%. Using the chronic pH
relationship, 64 mg N/L at pH=7.4 corresponds to 33 mg N/L at pH=8. The LC20
can be estimated to be 28 mg N/L based on the average ratio (0.82) between an
LC20 and LC30 in the fathead minnow chronic tests of DeGraeve et al. Regarding
duration, the heaviest mortality in this study was in the last few days, so LC20s at
even slightly longer durations would be less, but it is uncertain by how much. Even
if the 30-day LC20 is just 25% less than the 21-day LC20, when combined with the
factor of 1.5 used above for adjusting 30-day tests to long term mortality, the long-
term chronic value would be estimated to be no higher than 14 mg N/L.
Diamond et al. (1993) also conducted a 14-day exposure at 13°C and for this
exposure they reported 80% mortality at 53 mg N/L total ammonia (pH=7.7), but
none at 21 mg N/L and below. Linear interpolation of mortality versus log
concentration would suggest an LC20 of 29 mg N/L. Adjustment to pH=8 would
lower this value to 20 mg N/L. This should be lowered by at least a factor of 2 to
71
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account for the short duration of this test, suggesting the chronic value should be
no higher than 10 mg N/L. However, this test is made uncertain by a variety of
factors. First, the mortality was mainly in the first four days so that the effect
concentration is actually mainly determined by acute mortality, although some
mortality at 53 mg N/L occurred throughout the two weeks. Increased temperatures
and decreased pHs occurred on the fifth through seventh days. Dissolved oxygen
concentrations were not measured for the first five days of exposure and were low
on the sixth day, although they were acceptably high through the rest of the
exposure.
In the experimental stream study of Hermanutz et al. (1987), bluegills with an initial
average weight of 2.2 g grew during a 90-day exposure to an average of 26.9 g,
28.1 g, and 26.0 g in the control and lowest two exposure concentrations,
respectively. At the highest exposure concentration (9.4 mg N/L total ammonia,
pH=8.2, T=21 °C), the average final weight was 14.6 g, about a 50% reduction in
growth. Adjusted to pH=8, this corresponds to an EC50 of 12.7 mg N/L. This
would be reduced to a value of 7.6 by applying an EC20:EC50 ratio (=0.6) from the
channel catfish chronic growth study of Colt and Tchobanogous. Again, while this
study might not be applied directly to criterion calculations, it does support the
need for a GMCV<10 mg N/L for older bluegills.
Unlike for fathead minnow and channel catfish, this information does not allow
good quantification of a chronic effect concentration for non-early-life-stage
bluegills. The concentration for bluegill will be quantified using procedures
proposed in the next subsection.
Other fish
Although chronic ammonia toxicity to older life stages of other warm water species
have not been tested in the laboratory, chronic tests with juvenile salmonids (as
summarized previously) also suggest that GMCVs should be below 15 mg N/L (at
pH=8). If the limited duration of these tests is considered, concentrations should
probably be restricted to below 10 mg N/L. In the experimental stream study of
Hermanutz et al. (1987), rainbow trout showed significantly reduced survival and/or
growth in three separate exposures at mean total ammonia concentrations of from
4.4 to 8.0 mg N/L (adjusted to pH=8). Also in this stream study, in two six-week
exposures using walleye pike, complete mortality was observed at concentrations
of 14 to 21 mg N/L and reduced growth was observed at 3.8-7.0 mg N/L (adjusted
to pH=8).
While such data are not directly used in criteria derivation, they do support an
expectation that juvenile fish chronic values should be <10 mg N/L. But this
doesn't allow actual computation of "non-early-life-stage" (non-ELS) GMCVs for
warm water fish species for which there are no juvenile chronic tests.
To quantify this concentration, one approach would be to apply the ratios among
endpoints discussed above for fathead minnow and channel catfish. However, the
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ratio for the non-ELS to ELS chronic concentrations is variable: about 1 for the
catfish, 3 for the fathead minnow and maybe even higher given the information
presented above on the bluegill. The ratio for the SMAV to non-ELS chronic
concentration is more constant: 4.4 for the fathead and 5.5 for the catfish. This
latter ratio is analogous to the acute-chronic ratios often used in criteria, except
estimated here using non-ELS chronic effects. However, the average of the above
two values (approximately 5) is too high to apply to some other species of concern,
because it would place the non-ELS chronic values at or below the ELS chronic
values. Furthermore, this approach does not take into account how close acute
and chronic toxic effects concentrations are for a particular species.
Another approach to quantifying this concentration is to note that, for both fathead
minnow and channel catfish, the non-ELS chronic effect concentration is closer, on
a relative scale, to the early-life-stage effect concentration than it is to the SMAV
for these species (i.e., the ratio of non-ELS to ELS chronic is less than the ratio of
acute to non-ELS chronic). In the absence of good non-ELS information, a
reasonable default assumption would be to set the non-ELS SMCV as the
geometric mean of the ELS SMCV and the SMAV. This makes the non-ELS
concentration have the same ratio to both of the other concentrations.
Using this procedure, the non-ELS GMCVs will be 8.78 mg N/L for Lepomis and
9.55 mg N/L for Micropterus. For Lepomis, this number is consistent with the
information presented above which suggested concentrations below 10 mg N/L
might be needed. It is several times the SMCV for bluegill and 50% greater than
the SMCV for green sunfish. It is about three-fold lower than the SMAV for both
species, making it a reasonable factor below the acute concentrations. For
Micropterus, this number is slightly more than double the SMCV and slightly less
than half the SMAV.
Endpoint Selection for Invertebrates
Two of the four most sensitive GMCVs used to quantify the chronic criterion were for
invertebrates: the fingernail clam Musculium and the amphipod Hyalella. The endpoint
for the fingernail clam toxicity tests was 6-week survival of juvenile clams, and would be
relevant to all temperatures. The endpoint for the amphipod toxicity tests was
reproduction, which would not be relevant for low temperatures. However, the LC20 for
juvenile survival in these ten-week tests was the same as the effects concentration
used for reproduction, the difference being that for reproduction this concentration
actually was for >20% effect and was reported as a "less than" value, because the IC20
could not be reliably calculated. In any event, the same concentration can be used for
lower temperatures because it applies to juvenile and adult mortality as well as to
reproduction.
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Formulation of the CCC
Nine Genus Mean Chronic Values (GMCVs) are presented in Table 5. Although
Table 5 contains chronic data for the genus Oncorhynchus, no GMCV is derived
because of the large range in the EC20s; rather these chronic data will be used to
evaluate whether the CCC poses a risk to this genus.
Although Table 5 does not contain data for an insect genus, available information
concerning a stonefly (Thurston et al. 1984a) indicates that at least one species is
relatively resistant to ammonia. Therefore, calculation of the fifth percentile directly
from the GMCVs in Table 5 should adequately reflect the intent of the 1985 Guidelines.
For this calculation the number of tested species was taken to be 10, because a GMCV
for an insect is assumed to be greater than the four lowest GMCVs. The fifth percentile
value calculated by this procedure could be considered to be a "less than" value
because the lowest two GMCVs are "less than" values.
The relevant relationships for formulating a seasonal, pH- and temperature-dependent
chronic criterion are summarized below.
pH-dependence of chronic toxlclty
The chronic pH dependence derived from tests with Micropterus and Daphnia are
used here. The acute pH dependence is not applied to chronic toxicity because
the measured acute-chronic ratios change substantially with pH. Equation 12,
presented earlier in the document, describes the shape of chronic pH dependence.
Temperature-dependence of chronic toxicity - fish
Available data suggest minimal dependence offish ammonia toxicity on
temperature. Even where some acute toxicity data indicate higher LCSOs at low
temperatures (i.e., for fathead minnows), applying this to chronic toxicity is
contraindicated because available data, as well as theoretical considerations,
suggest that acute-chronic ratios increase at lower temperatures. Although limited
available chronic data suggest LC20s might be lower at low temperatures, the
effect is small and uncertain. This criteria formulation assumes no temperature
dependence for fish endpoints.
Temperature-dependence of chronic toxicity - invertebrates
Available data suggest a strong dependence of invertebrate acute ammonia toxicity
on temperature, with a mean log LC50 versus temperature slope of -0.036 over the
entire temperature range, or slightly higher, around -0.044 above approximately
10°C. No data are available regarding the temperature dependence of
invertebrate chronic toxicity, but both the fish data and theoretical expectations
argue against assuming that the chronic slope is as steep as the acute slope.
Alternative approaches for the chronic slope were discussed at the end of the
chronic temperature relationship section earlier in the document. The approach
selected for calculating the criterion is to set the invertebrate chronic slope equal to
74
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an invertebrate acute slope of -0.044 above 7°C, minus the observed fish ACR
slope of -0.016. This yields an invertebrate chronic slope of -0.028 above 7°C.
The slope is then set to zero below 7°C.
Endpoint selection - fish
The GMCVs appropriate for fish depend on whether early life stages of fish are
present. Determining when or whether time periods exist when such life stages are
present requires some information about the site or eco-region. The national
criterion will specify two sets of numbers between which the applied criterion can
select depending on the presence or absence offish early life stages.
When fish early life stages are present, the GMCVs are as presented in Table 5,
that is, 2.85 mg N/L for Lepomis, 3.09 mg N/L for Pimephales, 4.56 mg N/L for
Micropterus, and 8.84 mg N/L for Ictalurus). When fish early life stages are not
present, the assigned GMCVs are 9.30 mg N/L for Pimephales, 8.84 mg N/L for
Ictalurus, 8.78 mg N/L for Lepomis, and 9.55 mg N/L for Micropterus. Although
these GMCVs are uncertain, the calculation of the criterion ends up being
insensitive to these uncertainties. That is, because invertebrate GMCVs control
the value of the criterion, the fish non-ELS GMCVs would need to be reduced by
nearly 50% before they would affect the criterion.
Endpoint selection - invertebrates
The GMCV for Musculium is applicable to all temperatures because it is for long-
term juvenile survival. The GMCV for Hyalella is applicable to all temperatures
because, although it was derived for reproductive effects, the same effects
concentration would be obtained based on juvenile survival. Therefore, the
temperature-normalized GMCVs in Table 5 are used: 2.26 mg N/L for Musculium
and 1.45 mg N/L for Hyalella.
Temperature- and pH-dependent Criteria Calculation
Part of a criterion derivation is the estimation of the fifth percentile of sensitivity based
on the set of toxicity values available for different genera. This fifth percentile estimate
is intended to be what would be obtained by simple inspection if many genera had
been tested. When the number of tested genera is less than 19, this will ordinarily
entail an extrapolation below the lowest value, because, in such small sets, the lowest
value would not be expected to be as low as the fifth percentile (i.e., if many more
genera were tested, it is likely that the fifth percentile genus would be more sensitive
than any in the small data set). One characteristic of such extrapolations is that, if the
genera vary widely in sensitivity, the extrapolated criterion value will be further below
the lowest value than if the genera are tightly grouped. This is statistically appropriate
because if the variance is high, the fifth percentile of a distribution would be expected
to lie further below the lowest datum of a small data set than if the variance was low.
75
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However, while this characteristic of extrapolations is desirable for criteria derivations
across chemicals with different variances for genus sensitivities, it is not necessarily
appropriate when the genera are following different temperature or seasonal
dependencies, as is the case here. As sensitivities change with temperature or life
stage, the spread of the lowest four GMCVs changes, and thus the degree of
extrapolation also changes, such that the criterion does not end up increasing in
concert with reduced sensitivities as the temperatures decrease. That is, if the fifth
percentile genus value is re-extrapolated at each different temperature, the resulting
temperature dependence of the criterion would be found not to accord with any
temperature dependence that could be reasonably assumed for untested sensitive
species. Therefore, while temperature dependent criteria could be calculated by
developing sets of GMCVs for each temperature and computing the CCC from the four
most sensitive GMCVs at each temperature, this approach will not be used here.
Rather, the derivation here will start with a chronic criterion calculated at 25°C. This
reference temperature was selected because it is the temperature for the chronic test
with the most sensitive genus (Hyalella), as well as being within 3°C of all the other
tests with the four most sensitive genera. Consequently, temperature extrapolation
uncertainties are minimized. Table 5 presented the GMCVs normalized to 25°C. The
GMCVs for the four most sensitive species are thus as follows:
1.45 mg N/L for Hyalella
2.26 mg N/L for Musculium
2.85 mg N/L for Lepomis
3.09 mg N/L for Pimephales.
With N=10, this results in a CCC of 1.24 mg N/L at 25°C and pH=8. Figure 13 shows
the ranked GMCVs and the CCC, all at pH=8 and 25°C.
At 25°C, the 1999 CCC is two percent lower than the 1998 CCC, due to the
temperature adjustment of the Musculium EC20s from the tested temperatures to the
25°C reference temperature. (In contrast to the 1999 CCC, the 1998 CCC did not
change with temperature.)
The CCC of 1.24 mg/L is 15 percent lower than the lowest GMCV (1.45 mg N/L for
Hyalella). This degree of extrapolation below the lowest GMCV is modest and
reasonable given the low number of tested genera, and will be used for criteria
calculations for all other temperatures and for conditions offish early life stages
present or absent. As discussed above, the alternative of calculating the CCC directly
from new sets of GMCVs for each condition results in different degrees of extrapolation,
ranging up to 50% below the lowest GMCV, an extrapolation that would not be
reasonable here.
Because the most sensitive genera are invertebrates, the criterion will vary with
temperature according to the invertebrate chronic temperature relationship, but cannot
exceed 85.4% of the lowest fish GMCV. The lowest seasonal GMCVs for fish are
76
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2.85 mg N/L for Lepomis early life stages, and 8.78 mg N/L for Lepomis juveniles and
adults. When fish early life stages are present, the upper limit would be 2.43 mg N/L
(that is, 0.854 • 2.85 mg N/L lowest fish GMCV). When fish early life stages are
absent, the upper limit (0.854 • 8.78 mg N/L lowest juvenile and adult fish GMCV) is
never reached under any temperature condition. Therefore, the CCC equals 85.4
percent of lower of (a) the temperature-adjusted Hyalella GMCV, or (b) the lowest fish
GMCV. With fish early life stages present, at pH=8 the criterion would be expressed
as:
CCC = 0.854 • MIN(2.85,
(19)
This function increases steadily with decreasing temperature, T, until it reaches its
maximum (0.854 • 2.85) at 14.5°C, below which it remains constant.
Figure 13. Ranked Genus Mean Chronic Values (GMCVs) with the
Criterion Continuous Concentration (CCC).
100
oo
ii
Q.
10--
o
^
o
CCC
Ranked Genera
77
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With early life stages absent, at pH=8 the 2.85 mg N/L GMCV for early life stages of
Lepomis would have been replaced by the 8.78 mg N/L GMCV for juvenile and adult
Lepomis. However, since the latter GMCV is so high that it is never limiting, it can be
dropped. In this case the invertebrate slope plateau will become limiting below 7°C,
and this condition is therefore incorporated into the equation for early life stages absent
(at pH=8):
CCC = 0.854 • 1.45 • ioao28'(25-MAX(T>7)) (20)
This function increases steadily with decreasing temperature, T, until it reaches its
plateau at 7°C.
The pH dependency is incorporated through Equation 12, which converts chronic effect
concentrations from pH=8 to any other pH. Consequently, the seasonally varying, pH-
and temperature-dependent CCC would be expressed as follows.
For fish early life stages present:
CCC = 0.854 • °-0676 + - - - . M|N( 2.85 , 1.45 -10° 028'<25T>)
+ 1 07.688-pH 1 +10PH -7.688, '
For fish early life stages absent,
CCC = 0 854 • f °-0676 + _ 2'912 _ ) • 1 45 • 1 oao28'(25-MAX(T>7»
^ 1 + 1 07.688-pH 1 + 1gpH-7.688,/
In the final criteria statement, these equations can be shortened by multiplying
coefficients together.
The 1999 CCC with fish early life stages present is plotted as the "new" CCC in
Figure 14, along with the 1984/1985 "old" CCC and the EC20s from Table 5. The 1999
CCC is near the 1984/1985 CCC in the range of pH from about 7.5 to 8, but is
increasingly higher than the old CCC at lower and higher values of pH. At pH=8 and
25°C, the new CCC corresponds to acute-chronic ratios of (14.4 mg N/L)/(1.24 mg N/L)
= 11.6 using the calculated FAV when salmonids are present (but not lowered to
protect large rainbow trout) and (16.8 mg N/L)/(1.24 mg N/L) = 13.5 using the FAV
when salmonids are absent. These are in the range of the ACRs that can be derived
from the EC20s in Table 5 (see Appendix 7). The ACR used to calculate the old CCC
was 13.5 (Heber and Ballentine 1992).
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Figure 14. Chronic EC20s used in criteria derivation in relationship to Criterion
Continuous Concentrations (CCCs). The curve labeled "New CCC" is the
1999 CCC at a temperature of 24.6°C, at which it equals the 1998 CCC.
The "Old CCC" is from 1984/1985.
O)
E
.
O
50
20
10
5 1
0.5
0.2
0.1
O
O
»_o»
New CCC
—^
Old CCC, 10 C
Old CCC, 20 C
O Musculium
D Ceriodaphnia
O Daphnia
O Hyalella
• Pimephales
A Catostomus
V Ictalurus
• Oncorhynchus
+ Lepomis
• Micropterus
I
7
8
PH
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Several points should be noted concerning the CCC:
a. The two lowest GMCVs are "less than" values. The CCC would be lower if a point
estimate, rather than a "less than" value, could have been derived from the
Borgmann (1994) study with Hyalella, the most sensitive genus. The CCC also
might be lower if a point estimate, rather than a "less than" value, could have been
derived from the studies with the fingernail clam.
b. At 25°C and pH=8 any substantial increase in the CCC derived with the procedures
in this 1999 Update would require a higher GMCV for Hyalella and possibly a
higher SMCV for the recreationally important bluegill.
c. Because acutely resistant taxa are under-represented in the chronic dataset in
Table 5, it could be argued that N, the number of genera used in the calculation of
the CCC, should be increased from 10 to a higher value. A reasonable increase in
N would not have a large effect, however. For example, adding three resistant
genera would raise the CCC less than 10 percent.
d. The central tendency of the available chronic EC20s for salmonids, even though
not used directly in the calculation of the CCC, indicates that these species would
probably be protected by the CCC. Nevertheless, in some tests, effects were
observed at concentrations below the lower-temperature values of the CCC. The
data suggest that there might be important differences between strains of rainbow
trout.
e. Some of the laboratory and field data for the fingernail clam, which might be
considered to have special ecological importance at some sites, indicate that this
species would be affected at concentrations below the CCC. Other data indicate
that it would not be affected by such concentrations. At most sites the intermittency
of exposures would probably reduce risk.
f. When a threatened or endangered species occurs at a site and sufficient data
indicate that it is sensitive at concentrations below the CCC, it is appropriate to
consider deriving a site-specific criterion.
g. Partly for statistical reasons, the CCC is based on a 20 percent reduction in
survival, growth, and/or reproduction. Whether the maximum acceptable percent
reduction should be lower or higher than 20 percent under a set of conditions is a
risk management decision. Consult Appendix 6 for regression parameters to
calculate ECs corresponding to other percentages.
h. If it had been derived using available acute-chronic ratios (see Appendix 7), the
CCC would be greater than 2 mg N/L, which would be inappropriate because (1) it
would be above one of the GMCVs in a dataset for which N is only 10, (2) it would
not appear to protect early life stages of the recreationally important bluegill, and
(3) it might not protect the fingernail clam.
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Chronic Averaging Period
The averaging period for a CCC often needs to be shorter than the length of the tests
upon which it is based for two main reasons. First, concentrations in the field are
typically much more variable than concentrations in laboratory tests, and variable
concentrations of ammonia have been shown to be more toxic than constant
concentrations when the comparisons are based on average concentrations during the
exposure (Thurston et al. 1981 a). By shortening the averaging period to which the
CCC applies, the average concentration over the entire exposure will be below the
CCC, increasingly so as the variability of the concentration increases. Second, chronic
tests generally encompass different life stages, which might have different sensitivities,
so that effects might be elicited only, or disproportionately, during the fraction of the
test in which a sensitive life stage is present, rather than cumulatively over the whole
test. The 1984/1985 ammonia criteria document specified a CCC averaging period of 4
days as recommended in the 1985 Guidelines (U.S. EPA 1985b), except that an
averaging period of 30 days could be used when exposure concentrations were shown
to have "limited variability". The purpose of this section is to better define when a 30-
day averaging period is acceptable.
Tests having different durations and/or starting with organisms of different ages can
indicate how restrictive the averaging period needs to be. The best information
available is for the fathead minnow. Based on 7-day tests, EC20s of 7.08 mg N/L at
pH=8.34 and 5.25 mg N/L at pH=8.42 were calculated from the data of Willingham
(1987) and CVs of 8.37 mg N/L at pH=8 and 3.87 mg N/L at pH=8.5 were reported by
Camp Dresser and McKee (1997). Adjusted to pH=8, these concentrations are 12.1,
10.25, 8.37, and 8.65 mg N/L, respectively, with a geometric mean of 9.7 mg N/L. This
is approximately 2.5 times the geometric mean EC20 for the 30-day early life-stage
tests conducted by Swigert and Spacie (1983) and Mayes et al. (1986) as discussed
above. This suggests that the CCC averaging period could be 30 days, as long as
excursions above the CCC are restricted sufficiently to not exceed the mean EC20 from
the 7-day tests. A rigorous definition of this excursion restriction is not possible with
the limited data available, especially because no information is available concerning
the effects of variations within the 7-day period. It is convenient, however, to base the
excursion restriction on a 4-day period, because this period is the default that already
has to be considered in calculations of water quality-based effluent limits, and because
it provides a substantial limitation of variability relative to the 7-day EC20s. It is
uncertain how much higher than the CCC the 4-day average can be, but based on
these fathead minnow test results, 2.5-fold higher concentrations should be acceptable.
Some other data support the use of a longer averaging period. For example, the
studies of Anderson et al. (1978) and Sparks and Sandusky (1981) with fingernail
clams showed that effects gradually accumulated during exposures, suggesting that
longer averaging periods are acceptable. Also, in the field study at Monticello, time
variations in pH yielded time variations in the applicable CCC. Analysis of the data
presented by Zischke and Arthur (1987) for the fingernail clam indicated that limiting
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the highest 4-day average concentration to 2.0-2.5 times the CCC would protect this
species, whereas application of a 30-day average without this stipulation would allow
substantial effects on this species. In addition, Calamari et al. (1977,1981) and Solbe
and Shurben (1989) found that longer exposures of embryos and fry resulted in much
lower LCSOs than 96-hr exposures.
In contrast, some other studies suggest possible risks from longer averaging periods
under variable concentrations. For channel catfish, Bader (1990) reported a 24%
reduction in growth at 2.4 mg N/L in 7-day tests with young fry at pH=8.2; this
corresponds to just 3.3 mg N/L at pH=8, which is lower than the adjusted EC20s
reported from longer early-life stage tests and juvenile tests in Table 5. This suggests
that a short averaging period is advisable, but such a conclusion is very uncertain
because it involves interlaboratory comparisons with very few data and because Bader
(1990) also found similar sensitivity with older fry, so his results might represent a high
sensitivity of the test stock rather than factors relevant to the averaging period. A short
averaging period might also be inferred by the fact that the fathead minnow life-cycle
test (Thurston et al. 1986) showed an EC20 of 2.0 mg N/L for embryo hatchability,
substantially lower than for early life-stage tests. It is possible that this greater
sensitivity might be due to exposures starting earlier in the life-cycle tests than in the
early life-stage tests. The importance of early exposure to embryos was demonstrated
by Solbe and Shurben (1989) for rainbow trout. However, they dealt with a one-week
delay in exposures rather than <1 day and there are other possibilities for the more
sensitive results of Thurston et al. (1986).
Based on the fathead minnow early life-stage data, a 30-day averaging period is
justified with the restriction that the highest 4-day average within the 30 days is no
greater than twice the CCC. The data of Bader (1990) and Thurston et al. (1986)
suggest a potential risk from long averaging periods during fish spawning season, but
the evidence is weak and, even if variability within long averaging periods produces
short exposures that are sufficiently high to affect young embryos, only a small fraction
of total reproduction would generally be affected. A high priority should be given to
research to resolve how to better address different time-series of exposure.
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The National Criterion For Ammonia in Fresh Water
The available data for ammonia, evaluated using the procedures described in the
"Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of
Aquatic Organisms and Their Uses", indicate that, except possibly where an unusually
sensitive species is important at a site, freshwater aquatic life should be protected if
both of the following conditions are satisfied for the temperature, T, and pH of the
waterbody:
1. The one-hour average concentration of total ammonia nitrogen (in mg N/L) does
not exceed, more than once every three years on the average, the CMC (acute
criterion) calculated using the following equations. Where salmonid fish are
present:
rMr 0.275 39.0
CMC = +
-7.204
Or where salmonid fish are not present:
0.411 58.4
CMC =
-7.204
2A. The thirty-day average concentration of total ammonia nitrogen (in mg N/L) does
not exceed, more than once every three years on the average, the CCC (chronic
criterion) calculated using the following equations.
When fish early life stages are present:
CCC = [ °-0577 + ^±87 \ . M|N(285 145.10o.o28.(25 T))
^ 1 +1g7.688-pH 1 + 1gpH-7.688,/
When fish early life stages are absent:
QQQ = I 0-0577 + 2.487 I _ ^ ^ _ <|Q0.028-(25-MAX(T,7))
-I +-|Q7.688-pH -I +<|QpH- 7.688
2B. In addition, the highest four-day average within the 30-day period should not
exceed 2.5 times the CCC.
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Several points should be noted concerning the criterion:
1. The two lowest GMCVs are "less than" values. The CCC would be lower if a point
estimate, rather than a "less than" value, could have been derived from the
Borgmann (1994) study with Hyalella, the most sensitive genus. The CCC also
might be lower if a point estimate, rather than a "less than" value, could have been
derived from the studies with the fingernail clam.
2. The central tendency of the available chronic EC20s for salmonids, even though
not used directly in the calculation of the CCC, indicate that these species would
probably be protected by the CCC, although the data suggest that there might be
important differences between strains of rainbow trout, and some tests indicated
effects at concentrations below the CCC.
3. Some of the laboratory and field data for the fingernail clam, which might be
considered to have special ecological importance at some sites, indicate that this
species would be affected at concentrations below the CCC. Other data indicate
that it would not be affected by such concentrations. At most sites the intermittency
of exposures would probably reduce risk.
4. When a threatened or endangered species occurs at a site and sufficient data
indicate that it is sensitive at concentrations below the CCC, it is appropriate to
consider deriving a site-specific criterion.
5. Partly for statistical reasons, the CCC is based on a 20 percent reduction in
survival, growth, and/or reproduction. Whether the maximum acceptable percent
reduction should be lower or higher than 20 percent under a set of conditions is a
risk management decision. ECs corresponding to other percentage reductions can
be calculated using the parameter values presented in Appendix 6.
The Recalculation Procedure, the WER Procedure, and the Resident Species
Procedure may be used to derive site-specific criteria for ammonia, but most WERs
that have been determined for ammonia are close to 1.
The CMC, CCC, and chronic averaging period presented above supersede those given
in previous guidance on the aquatic life criterion for ammonia in fresh water. The 1998
and 1999 Updates do not address or alter the past recommendation of a one-hour
averaging period for the CMC or the past recommendation of a once-in-three years on
the average allowable frequency for exceeding the CMC or CCC. Many issues
concerning the implementation of aquatic life criteria are discussed in the "Technical
Support Document for Water Quality-based Toxics Control" (U.S. EPA 1991).
If concentrations in 95 percent of grab or 24-hour composite samples do not exceed the
CCC, then the 30-day average concentrations are unlikely to exceed the CCC more
then once in three years (Delos 1999). This assumes that concentrations are log
normally distributed, with first-order log serial correlation coefficient between 24-hour
composite samples approximately 0.86 - 0.94 (or less), and a log standard deviation of
0.5-0.8 (or less).
Because the ammonia criterion is a function of pH and temperature, calculation of the
appropriate weighted average temperature or pH is complicated. For some purposes,
84
-------
calculation of an average pH and temperature can be avoided. For example, if
samples are obtained from a receiving water over a period of time during which pH
and/or temperature is not constant, the pH, temperature, and the concentration of total
ammonia in each sample should be determined. For each sample, the criterion should
be determined at the pH and temperature of the sample, and then the concentration of
total ammonia nitrogen in the sample should be divided by the criterion to determine a
quotient. The criterion is attained if the mean of the quotients is less than 1 over the
duration of the averaging period.
85
-------
pH-Dependent Values of the CMC (Acute Criterion)
CMC,
PH
6.5
6.6
6.7
6.8
6.9
7.0
7.1
7.2
7.3
7.4
7.5
7.6
7.7
7.8
7.9
8.0
8.1
8.2
8.3
8.4
8.5
8.6
8.7
8.8
8.9
9.0
Salmonids
Present
32.6
31.3
29.8
28.1
26.2
24.1
22.0
19.7
17.5
15.4
13.3
11.4
9.65
8.11
6.77
5.62
4.64
3.83
3.15
2.59
2.14
1.77
1.47
1.23
1.04
0.885
mg N/L
Salmonids
Absent
48.8
46.8
44.6
42.0
39.1
36.1
32.8
29.5
26.2
23.0
19.9
17.0
14.4
12.1
10.1
8.40
6.95
5.72
4.71
3.88
3.20
2.65
2.20
1.84
1.56
1.32
86
-------
Temperature and pH-Dependent Values of the CCC (Chronic Criterion)
for Fish Early Life Stages Present
CCC for
PH
6
6
6
6
6
7
7
7
7
7
7
7
7
7
7
8
8
8
8
8
8
8
8
8
8
9
5
6
7
8
9
0
1
2
3
4
5
6
7
8
9
0
1
2
3
4
5
6
7
8
9
0
0
6.67
6.57
6.44
6.29
6.12
5.91
5.67
5.39
5.08
4.73
4.36
3.98
3.58
3.18
2.80
2.43
2.10
1.79
1.52
1.29
1.09
0.920
0.778
0.661
0.565
0.486
14
6.67
6.57
6.44
6.29
6.12
5.91
5.67
5.39
5.08
4.73
4.36
3.98
3.58
3.18
2.80
2.43
2.10
1.79
1.52
1.29
1.09
0.920
0.778
0.661
0.565
0.486
16
6.06
5.97
5.86
5.72
5.56
5.37
5.15
4.90
4.61
4.30
3.97
3.61
3.25
2.89
2.54
2.21
1.91
1.63
1.39
1.17
0.990
0.836
0.707
0.601
0.513
0.442
Fish Early Life
18
5.33
5.25
5.15
5.03
4.89
4.72
4.53
4.31
4.06
3.78
3.49
3.18
2.86
2.54
2.24
1.94
1.68
1.43
1.22
1.03
0.870
0.735
0.622
0.528
0.451
0.389
Stages Present,
Temperature, C
20 22
4.68
4.61
4.52
4.42
4.30
4.15
3.98
3.78
3.57
3.32
3.06
2.79
2.51
2.23
1.96
1.71
1.47
1.26
1.07
0.906
0.765
0.646
0.547
0.464
0.397
0.342
4.12
4.05
3.98
3.89
3.78
3.65
3.50
3.33
3.13
2.92
2.69
2.45
2.21
1.96
1.73
1.50
1.29
1.11
0.941
0.796
0.672
0.568
0.480
0.408
0.349
0.300
24
3.62
3.56
3.50
3.42
3.32
3.21
3.08
2.92
2.76
2.57
2.37
2.16
1.94
1.73
1.52
1.32
1.14
0.973
0.827
0.700
0.591
0.499
0.422
0.359
0.306
0.264
mg N/L
26
3.18
3.13
3.07
3.00
2.92
2.82
2.70
2.57
2.42
2.26
2.08
1.90
1.71
1.52
1.33
1.16
1.00
0.855
0.727
0.615
0.520
0.439
0.371
0.315
0.269
0.232
28
2.80
2.75
2.70
2.64
2.57
2.48
2.38
2.26
2.13
1.98
1.83
1.67
1.50
1.33
1.17
1.02
0.879
0.752
0.639
0.541
0.457
0.386
0.326
0.277
0.237
0.204
30
2.46
2.42
2.37
2.32
2.25
2.18
2.09
1.99
1.87
1.74
1.61
1.47
1.32
1.17
1.03
0.897
0.773
0.661
0.562
0.475
0.401
0.339
0.287
0.244
0.208
0.179
87
-------
Temperature and pH-Dependent Values of the CCC (Chronic Criterion)
for Fish Early Life Stages Absent
PH
6
6
6
6
6
7
7
7
7
7
7
7
7
7
7
8
8
8
8
8
8
8
8
8
8
9
5
6
7
8
9
0
1
2
3
4
5
6
7
8
9
0
1
2
3
4
5
6
7
8
9
0
0-7
10.8
10.7
10.5
10.2
9.93
9.60
9.20
8.75
8.24
7.69
7.09
6.46
5.81
5.17
4.54
3.95
3.41
2.91
2.47
2.09
1.77
1.49
1.26
1.07
0.917
0.790
8
10.1
9.99
9.81
9.58
9.31
9.00
8.63
8.20
7.73
7.21
6.64
6.05
5.45
4.84
4.26
3.70
3.19
2.73
2.32
1.96
1.66
1.40
1.18
1.01
0.860
0.740
CCC for
9
9.51
9.37
9.20
8.98
8.73
8.43
8.09
7.69
7.25
6.76
6.23
5.67
5.11
4.54
3.99
3.47
2.99
2.56
2.18
1.84
1.55
1.31
1.11
0.944
0.806
0.694
Fish Early Life
10
8.92
8.79
8.62
8.42
8.19
7.91
7.58
7.21
6.79
6.33
5.84
5.32
4.79
4.26
3.74
3.26
2.81
2.40
2.04
1.73
1.46
1.23
1.04
0.885
0.756
0.651
Stages
Temperature
11 12
8.36
8.24
8.08
7.90
7.68
7.41
7.11
6.76
6.37
5.94
5.48
4.99
4.49
3.99
3.51
3.05
2.63
2.25
1.91
1.62
1.37
1.15
0.976
0.829
0.709
0.610
7.84
7.72
7.58
7.40
7.20
6.95
6.67
6.34
5.97
5.57
5.13
4.68
4.21
3.74
3.29
2.86
2.47
2.11
1.79
1.52
1.28
1.08
0.915
0.778
0.664
0.572
Absent,
13
7.35
7.24
7.11
6.94
6.75
6.52
6.25
5.94
5.60
5.22
4.81
4.38
3.95
3.51
3.09
2.68
2.31
1.98
1.68
1.42
1.20
1.01
0.858
0.729
0.623
0.536
mg N/L
14
6.89
6.79
6.66
6.51
6.33
6.11
5.86
5.57
5.25
4.89
4.51
4.11
3.70
3.29
2.89
2.52
2.17
1.85
1.58
1.33
1.13
0.951
0.805
0.684
0.584
0.503
15*
6.46
6.36
6.25
6.10
5.93
5.73
5.49
5.22
4.92
4.59
4.23
3.85
3.47
3.09
2.71
2.36
2.03
1.74
1.48
1.25
1.06
0.892
0.754
0.641
0.548
0.471
16*
6.06
5.97
5.86
5.72
5.56
5.37
5.15
4.90
4.61
4.30
3.97
3.61
3.25
2.89
2.54
2.21
1.91
1.63
1.39
1.17
0.990
0.836
0.707
0.601
0.513
0.442
* At 15 C and above, the criterion for fish ELS absent is the same as the criterion for fish ELS
present.
88
-------
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Appendix 1. Review of Some Toxicity Tests
Diamond et al. (1993) reported results of a variety of acute and chronic toxicity tests on
ammonia. Data sheets and reports concerning these tests were examined for
additional information that would be useful in the evaluation of the tests and
interpretation of the results. The most common problem was that the concentration of
dissolved oxygen was too low or too high.
Water-Effect Ratios
The data sheets and reports revealed that the information in Table 2 in Diamond et al.
(1993) is correct. The invertebrate used was D. magna as stated on page 653, not D.
pulex as stated on page 652.
Acutetoxicitvat20°C
The data sheets and reports revealed the following regarding the information in Table
3:
a. The concentration of dissolved oxygen was above 110 percent of saturation for a
portion of the test with the bay silverside.
b. The highest tested concentration in the test with the bluegill killed only 40
percent of the test organisms.
c. The data sheets say that tests were conducted with two species of crayfish.
Subsequently, the authors said that it was later determined that Procambarus
clarkii was used in both tests and that all of the crayfish were obtained from the
same supplier. The LC50 in the table is from a test in which the concentration of
dissolved oxygen was below 44 percent of saturation for a portion of the test.
d. The LC50 given for the amphipod is a 21 -day LC50. The concentration of
dissolved oxygen was below 50 percent of saturation for a portion of the test.
e. The LC50 given for the spring peeper is a 9-day LC50.
Some of these tests were conducted in a laboratory dilution water and some were
conducted in a well water; these were the two waters used in the determination of the
Water-Effect Ratios (see above).
Chronic toxicitv at 20°C
The data sheets and reports revealed the following regarding the chronic tests that are
the basis of the results in Table 4:
Leopard frog (larvae-tadpole)
The concentration of dissolved oxygen was below 50 percent of saturation for a
portion of the test. In addition, this test lasted for only 14 days.
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Leopard frog (egg-larvae)
The concentration of dissolved oxygen was below 40 percent of saturation for a
portion of the test. In addition, this test lasted for only 20 days.
Bluegill
There were no major problems with this test, which was conducted in a
laboratory dilution water. The durations of the chronic tests with the bluegill in
warm and cold water are switched in Table 1.
Crayfish (Procambarus clarkii)
The concentration of dissolved oxygen was below 40 percent of saturation for a
portion of the test. In addition, this test lasted for only 21 days.
Amphipod (Crangonyx spp.)
The concentration of dissolved oxygen was below 40 percent of saturation for a
portion of the test. In addition, this test was begun with organisms that were 8 to
42 days old and lasted for only 21 days.
Acute toxicitv at 12°C
The data sheets and reports revealed the following concerning the information in Table
5:
a. The LC50 for the sheepshead minnow is a 48-hr LC50.
b. The data sheets say that the crayfish used was Astacus pallipes. Subsequently,
the authors said that it was later determined that the crayfish used was
Procambarus clarkii.
Some of these tests were conducted in a laboratory dilution water and some were
conducted in a well water; these were the two waters used in the determination of the
Water-Effect Ratios (see above).
Effect of temperature on the toxicitv of ammonia
The data sheets, reports, and publication revealed the following concerning the acute
values in Table 6:
1. A comparison is not possible for the dragonfly because both of the values are
"greater than" values.
2. The two acute tests with the bluegill were conducted in different waters.
3. One of the chronic tests with the bluegill lasted for 14 days, whereas the other
lasted for 21 days. The concentration of dissolved oxygen was below 40 percent
of saturation for a portion of the 14-day test.
4. For the amphipod, the LC50 at 12°C is a 96-hr LC50, whereas the LC50 at 20°C
is a 21 -day LC50. In the 21 -day test, the concentration of dissolved oxygen was
below 50 percent of saturation for a portion of the test.
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5. The two tests with crayfish were conducted in different waters. In the test at
20°C, the concentration of dissolved oxygen was below 40 percent of saturation
for a portion of the test. The LC50 at 12°C was ">2.35" as reported in Table 5,
not "2.35" as reported in Table 6.
6. The NOEC of 0.44 mg/L given in Table 6 for the leopard frog at 12°C is from a
test with the spring peeper.
7. The concentration of dissolved oxygen was above 110 percent of saturation for a
portion of one of the tests with the bay silverside.
8. The LC50 given in Table 6 for the spring peeper at 20°C is a 9-day LC50,
whereas the value at 12°C is a 96-hr LC50. Because the 9-day LC50 at 20°C is
greater than the 96-hr LC50 at 12°C, a qualitative comparison is possible.
Valid comparisons of 12 versus 20°C can be made only for the two amphibians.
The data sheets, reports, and publication revealed the following concerning the chronic
tests that are the basis of the results in Table 6:
The three chronic tests at 20°C were addressed above.
Bluegillat 12°C:
The concentration of dissolved oxygen was below 40 percent of saturation for a
portion of the test. In addition, this test was begun with juveniles and lasted for
only 14 days. (The durations of the chronic tests with the bluegill in warm and
cold water are switched in Table 1.)
The chronic comparison with the bluegill is based on a 21-day test and a 14-day
test. In addition, the concentration of dissolved oxygen was below 40 percent of
saturation during a portion of the test at 12°C.
Amphipod (Crangonyx spp.) at 12°C:
The concentration of dissolved oxygen was below 40 percent of saturation for a
portion of the test. In addition, this test was begun with juveniles and lasted for
only 21 days. In both of the chronic tests used in the chronic comparison with
the amphipod, the concentration of dissolved oxygen was below 40 percent of
saturation during a portion of the test.
Leopard frog at 12°C:
This chronic test was conducted with the spring peeper, not the leopard frog.
The concentration of dissolved oxygen was above 110 percent of saturation for a
portion of the test. In addition, this test was begun seven days after hatch and
lasted for only 21 days. The chronic comparison with the leopard frog is based
on a chronic test conducted with the leopard frog and a chronic test conducted
with the spring peeper.
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Appendix 2. Methods for Regression Analysis of pH Data
Analysis of the available data relating ammonia toxicity to pH using Equations 9 and 10
requires recognition that, unlike usual regression analysis with one response variable,
two response variables (i.e., LC50U and LC50,) are of concern here. Suitable analysis
requires some assumptions about the correlations among these response variables
(Box and Draper 1965; Box et al. 1973; Draper and Smith 1981). If the correlations
among the data are known, Box and Draper (1965) indicate that regression analysis
should involve minimization of the quantity:
z = E E °ijvy
(27)
u=1
where k is the number of dependent variables, n is the number of datapoints, yiu is the
observed value for the dependent variable i, and f(xiu,6) is the model prediction of the
value of the dependent variable i. If correlation coefficients are zero, Equation 27
reduces to standard least squares regression techniques. However, when correlations
are unknown, Box and Draper (1965) indicate that the determinant of the matrix of v^
should be minimized; this results in a formulation similar to Equation 27, but with
weights calculated from relationships within the data rather than from a priori
knowledge or assumptions regarding variances. If linear relationships exist among the
dependent variables, further refinements are necessary (Box et al. 1973). Before using
these more complicated techniques, which might have rather minimal impact on
parameter estimates, consideration was first given to what could be assumed about the
correlations of the errors in LC50U and LC50,.
Because LC50U and LC50, are both derived from LC50t based on chemical equilibrium
equations (i.e., Equation 4), it might be thought that their errors are directly correlated
and proportional to that of LC50t. However, uncertainty also exists in the equilibrium
fractions, mainly from uncertainty in pH, and this results in errors that are inversely
correlated. Lacking any definitive resolution of the degree of correlation, simulations
were run to determine whether methods assuming no correlation would produce
acceptable results. As mentioned above, this assumption results in applying standard
least squares regression techniques to Equations 9 and 10.
For assumed parameter values LC508=1.0, pHT=7.5, R=0.01, and o=0.1, four sets of
1000 simulations were run in which hypothetical datasets were randomly generated
and analyzed. The four sets differed based on a 2x2 arrangement of two factors, each
with two options. One factor was the size of the dataset - both small (n=5 with pH
ranging from 6.5 to 8.5 at 0.5 intervals) and large (n=13 with pH ranging from 6.0 to 9.0
104
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at 0.25 intervals) datasets were run. The other factor was the true correlation between
the errors in logl_C50u and logLCSOji one option had the correlation coefficient = 0
(which met analysis assumptions) and the other had the correlation coefficient = 1
(which violated analysis assumptions as much as possible). Estimates of the standard
errors of the parameters were based on the covariance matrix computed from the
residual error and inverse Jacobian at the least squares solution; confidence limits
were computed as the product of this standard error and the applicable t-statistic.
These simulations and their results are summarized in Table 6. Parameter values were
found to be unbiased in all cases. When true errors were uncorrelated, as assumed in
the procedure, the estimated parameter standard errors were unbiased relative to the
standard deviations of the estimated parameter values, and the confidence limits were
95% using 2n-3 degrees of freedom. When true errors were correlated, the estimated
parameter standard errors were biased, averaging 11 to 33% less than the observed
error in the estimated parameter values, and the confidence limits were 80 to 89%
rather than 95%. At the smallest sample size, the biases in the estimated errors were
only 0.05 units for pKT, 0.03 units for log10R (corresponding to 7% bias in the error for
R), and 0.01 units for log10LC50t8 (corresponding to only 2.5% bias in the error for
LC50t8). Because these biases were relatively small, because the actual parameter
estimates were unbiased, and because this analysis was under worst-case
assumptions, standard regression methods with the assumption of no correlation of
errors were adopted for the analysis of pH effects using Equations 9 and 10, rather
than adopting more complicated methods.
105
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Table 6. Results Obtained using Simulated Samples
Parameter
True Value
pKT
7.5
log10R
-2.0
log10LC50t8
0.0
Simulations with 5 Treatments - Errors Uncorrelated
Mean of Estimated Parameter Values
Standard Deviation of Estimated Parameter Values
Mean of Estimated Parameter Standard Errors
Simulated Confidence for Nominal 95% CL
7.501
0.104
0.104
95%
-1.994
0.123
0.121
95%
-0.001
0.050
0.051
96%
Simulations with 13 Treatments - Errors Uncorrelated
Mean of Estimated Parameter Values
Standard Deviation of Estimated Parameter Values
Mean of Estimated Parameter Standard Errors
Simulated Confidence for Nominal 95% CL
7.498
0.057
0.056
94%
-2.000
0.068
0.069
95%
-0.001
0.030
0.031
95%
Simulations with 5 Treatments - Errors Correlated
Mean of Estimated Parameter Values
Standard Deviation of Estimated Parameter Values
Mean of Estimated Parameter Standard Errors
Simulated Confidence for Nominal 95% CL
7.499
0.145
0.097
80%
-2.001
0.146
0.114
84%
0.003
0.058
0.047
86%
Simulations with 13 Treatments - Errors Correlated
Mean of Estimated Parameter Values
Standard Deviation of Estimated Parameter Values
Mean of Estimated Parameter Standard Errors
Simulated Confidence for Nominal 95% CL
7.501
0.079
0.055
82%
-1.999
0.079
0.067
89%
0.001
0.034
0.030
89%
106
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Appendix 3. Conversion of Results of Toxicity Tests
All of the acute values reported in Table 1 of the 1984/1985 ammonia criteria document
(U.S. EPA 1985a) are expressed in terms of un-ionized ammonia at the pH of the
toxicity test. For use in the 1998 and 1999 Updates, they were converted from un-
ionized ammonia at the test pH to total ammonia nitrogen at pH=8. The conversion
procedure is illustrated here using the data for the flatworm, Dendrocoelum lacteum,
which is the first species in Table 1 in the 1984/1985 criteria document and is the first
species in Appendix 4 in this 1999 Update:
Acute value (AV) = 1.40 mg NH3/L
pH = 8.20
Temperature = 18.0°C
Step 1.
Equation 3 in this 1999 Update is used to calculate the pK at 18°C:
pK = 9.464905
Step 2.
Equation 2 in this Update and the definitions pK = -log10K and pH = -log10[H+]
are used to obtain the following:
[NH,1 , u ^
= io(pH-pK) = 00543369
[NH4]
Step 3.
The AV in terms of total ammonia is calculated as:
Total ammonia = [NH3] + [NH4] = [NH3]
[NH3]
0.0543369
= 27.1652 mg total ammonia/L
Step 4.
The AV in terms of total ammonia nitrogen is calculated as follows:
Total ammonia nitrogen = (27.1652 mg total ammonia/L)(14/17)
= 22.3713 mgN/L
107
-------
Step 5.
The AV in terms of total ammonia nitrogen is converted from pH=8.2 to pH=8
using Equation 11 in this Update:
AVt8 = (AVt)/(0.681546) = 32.8244 mg N/L
Because this is the only species in this genus for which data are in Table 1 in the
1984/1985 criteria document, 32.82 mg N/L is the GMAV given for the genus
Dendrocoelum in Table 4 in this update.
108
-------
Appendix 4. Acute Values8
Species
Dendrocoelum lacteum
Tubifex tubifex
Physa gyrina
Physa gyrina
Physa gyrina
Physa gyrina
Physa gyrina
Physa gyrina
Helisoma trivolvis
Musculium transversum
Musculium transversum
Musculium transversum
Ceriodaphnia acanthina
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia pulicaria
Simocephalus vetulus
Simocephalus vetulus
Asellus racovitzai
Asellus racovitzai
Crangonyx pseudogracilis
Crangonyx pseudogracilis
Un-ionized
Ammonia
(mg NH3/L)
1.40
2.70
1.59
2.09
2.49
2.16
1.78
1.71
2.76
0.93
1.29
1.10
0.770
2.08
2.45
2.69
2.50
2.77
2.38
0.75
0.90
0.53
0.67
4.94
1.16
0.613
2.29
2.94
4.95
2.76
5.63
PH
8.20
8.20
8.00
8.20
8.10
8.20
8.00
8.00
8.20
8.20
8.10
8.60
7.06
8.20
7.95
8.07
8.09
8.15
8.04
7.51
7.53
7.40
7.50
8.34
8.05
7.06
8.30
7.81
8.00
8.00
8.00
Temp.
(°C)
18.0
12.0
4.0
5.5
12.1
12.8
13.3
24.9
12.9
5.4
14.6
20.5
24.0
25.0
22.0
19.6
20.9
22.0
22.8
20.1
20.1
20.6
20.3
19.7
14.0
24.0
17.0
11.9
4.0
4.0
12.1
Total
Ammonia
(mg N/L)
22.37
66.67
114.93
85.13
76.29
50.25
62.39
26.33
63.73
38.18
32.83
6.43
104.82
20.71
51.30
51.09
41.51
37.44
38.70
48.32
55.41
42.31
43.52
51.92
34.50
83.45
31.58
176.01
357.80
199.50
215.97
Total
Ammonia
(mg N/L@pH8)
32.82
97.82
114.87
124.90
92.27
73.73
62.36
26.32
93.52
56.02
39.70
20.38
25.78
30.38
46.68
58.33
49.25
49.86
41.73
20.72
24.49
15.48
18.39
100.02
37.91
20.52
56.29
124.02
357.60
199.39
215.85
Reference
Stammer 1953
Stammer 1953
West 1985
West 1985
West 1985
West 1985
West 1985
West 1985
West 1985
West 1985
West 1985
West 1985
Mount 1982
Parkhurst et al. 1979,1981
Russo etal. 1985
Russo etal. 1985
Russo etal. 1985
Russo etal. 1985
Russo etal. 1985
Russo etal. 1985
Russo etal. 1985
Russo etal. 1985
Russo etal. 1985
Reinbold & Pescitelli 1982a
DeGraeve etal. 1980
Mount 1982
West 1985
Thurston etal. 1983a
West 1985
West 1985
West 1985
109
-------
Crangonyx pseudogracilis
Crangonyx pseudogracilis
Crangonyx pseudogracilis
Orconectes nais
Orconectes immunis
Callibaetis sp.
Callibaetis skokianus
Ephemerella grandis
Ephemerella grandis
Ephemerella grandis
Arcynopteryx parallela
Arcynopteryx parallela
Philarctus quaeris
Stenelmis sexlineata
Oncorhynchus gorbuscha
Oncorhynchus gorbuscha
Oncorhynchus kisutch
Oncorhynchus kisutch
Oncorhynchus kisutch
Oncorhynchus kisutch
Oncorhynchus kisutch
Oncorhynchus kisutch
Oncorhynchus kisutch
Oncorhynchus kisutch
Oncorhynchus tshawytscha
Oncorhynchus tshawytscha
Oncorhynchus tshawytscha
Oncorhynchus aquabonita
Oncorhynchus clarki
Oncorhynchus clarki
Oncorhynchus clarki
Oncorhynchus clarki
3.56
3.29
1.63
3.15
22.8
1.80
4.82
4.96
5.88
3.86
2.06
2.00
10.2
8.00
0.083
0.10
0.272
0.280
0.550
0.528
0.712
0.700
0.880
0.55
0.476
0.456
0.399
0.755
0.80
0.66
0.62
0.52
8.20
8.00
8.00
8.30
8.20
7.81
7.90
7.84
7.85
7.84
7.76
7.81
7.80
8.70
6.40
6.40
7.00
7.00
7.50
7.50
8.00
8.00
8.50
8.10
7.82
7.84
7.87
8.06
7.81
7.80
7.80
7.78
13.0
13.3
24.9
26.5
4.6
11.9
13.3
12.8
12.0
13.2
13.8
13.1
13.3
25.0
4.3
4.30
15.0
15.0
15.0
15.0
15.0
15.0
15.0
17.2
12.2
12.3
13.5
13.2
13.1
12.8
12.4
12.2
81.60
115.32
25.10
23.15
999.39
107.76
21 1 .66
259.07
319.03
195.62
119.63
109.31
561 .72
29.69
230.47
277.68
82.02
84.43
52.76
50.65
22.00
21.63
9.09
11.59
27.23
24.74
18.47
23.30
43.72
37.75
36.55
32.57
119.73
115.25
25.08
41.27
1466.35
75.93
175.56
192.64
241 .54
145.46
77.18
77.03
388.84
113.17
38.33
46.18
19.10
19.66
22.29
21.40
21.99
21.62
23.86
14.02
19.53
18.39
14.50
26.10
30.81
26.13
25.30
21.76
West 1985
West 1985
West 1985
Evans 1 979
West 1985
Thurston et al. 1984a
West 1985
Thurston et al. 1984a
Thurston et al. 1984a
Thurston et al. 1984a
Thurston et al. 1984a
Thurston et al. 1984a
West 1985
Hazel etal. 1979
RiceS Bailey 1980
RiceS Bailey 1980
Robinson-Wilson & Seim
1975
Robinson-Wilson & Seim
1975
Robinson-Wilson & Seim
1975
Robinson-Wilson & Seim
1975
Robinson-Wilson & Seim
1975
Robinson-Wilson & Seim
1975
Robinson-Wilson & Seim
1975
Buckley 1978
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Russo 1981
Thurston et al. 1978
Thurston et al. 1978
Thurston et al. 1978
Thurston et al. 1978
110
-------
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
0.325
0.370
0.160
0.440
0.697
0.40
0.77
0.436
0.446
0.478
0.291
0.232
0.336
0.347
0.474
0.440
0.392
0.426
0.400
0.497
0.421
0.758
0.572
0.570
0.673
1.09
0.641
0.696
0.772
0.683
0.812
0.632
0.618
0.410
0.390
0.752
7.40
7.40
7.40
7.40
7.95
7.50
8.05
7.90
7.90
7.91
7.91
7.88
7.88
7.87
7.95
7.87
7.87
7.88
7.87
7.86
7.86
8.08
7.86
7.85
7.85
8.06
7.85
7.79
7.86
7.84
7.80
7.85
7.87
7.71
7.71
7.84
14.4
14.5
14.5
14.5
10.0
15.0
14.0
12.7
13.4
13.0
13.1
12.8
12.9
12.9
12.5
13.0
12.9
13.4
13.1
13.4
13.0
12.8
12.7
12.5
13.1
13.2
12.3
12.4
14.1
13.8
12.4
13.1
12.1
11.4
11.5
13.0
40.99
46.31
20.03
55.07
35.14
38.37
22.90
20.03
19.44
20.99
12.68
11.07
15.91
16.81
19.75
21.15
18.99
19.43
19.08
23.71
20.70
23.05
28.77
29.77
33.59
33.64
33.99
41.97
34.95
33.09
47.87
31.55
31.80
32.02
30.22
38.69
14.99
16.94
7.33
20.15
31.97
16.21
25.17
16.61
16.12
17.73
10.71
8.85
12.72
13.19
17.97
16.61
14.91
15.53
14.98
18.28
15.96
26.82
22.18
22.54
25.44
37.68
25.74
28.55
26.94
24.60
33.14
23.89
24.97
18.95
17.89
28.77
Calamari et al. 1977, 1981
Calamari et al. 1977, 1981
Calamari et al. 1977, 1981
Calamari et al. 1977, 1981
BroderiusS Smith 1979
Holt & Malcolm 1979
DeGraeve etal. 1980
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
111
-------
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
0.662
0.763
0.250
0.449
0.392
0.464
0.243
0.635
0.510
0.623
0.833
0.432
0.796
0.714
0.326
0.404
0.389
0.375
0.364
0.382
0.367
0.392
0.281
0.456
0.432
0.268
0.307
0.351
0.448
0.552
0.580
0.484
0.297
0.327
0.289
0.262
7.83
7.80
7.44
7.84
7.87
7.90
7.50
7.82
7.75
7.84
7.90
7.70
7.90
7.87
7.80
7.65
7.67
7.62
7.64
7.66
7.65
7.69
7.60
7.75
7.66
7.60
7.63
7.59
7.68
111
7.86
7.88
7.69
7.74
7.76
7.66
13.5
13.3
12.8
12.2
12.2
11.9
14.5
13.2
12.3
12.9
13.0
13.9
13.0
13.0
9.7
14.3
14.0
14.4
13.1
13.6
13.2
13.4
12.9
11.8
12.8
13.0
12.9
12.7
13.0
13.6
10.2
10.0
10.7
10.4
10.0
9.80
33.55
42.02
32.49
24.54
20.02
22.65
24.20
33.67
33.94
32.30
37.41
28.54
35.75
34.32
23.65
29.02
27.30
28.62
29.28
28.27
28.64
27.51
25.14
31.53
33.97
23.80
25.65
32.62
33.15
31.81
35.31
28.60
25.62
25.76
22.44
25.95
24.50
29.09
12.57
18.25
15.72
18.79
10.22
24.15
21.52
24.01
31.03
16.60
29.65
26.95
16.37
15.53
15.11
14.58
15.42
15.38
15.33
15.74
12.40
19.99
18.48
11.74
13.29
15.84
18.65
20.89
27.23
22.87
14.66
16.05
14.47
14.12
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
112
-------
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
0.312
0.201
0.234
0.249
0.192
0.163
0.677
0.662
0.636
0.694
0.764
0.921
0.856
0.801
0.897
0.942
0.931
0.158
0.184
0.454
0.799
0.684
0.648
0.683
0.704
0.564
0.610
0.497
0.643
0.56
0.79
0.40
1.02
0.77
0.97
0.26
7.64
7.69
7.69
7.64
7.65
7.62
8.10
8.12
7.94
7.98
7.89
7.94
7.85
7.88
7.91
7.91
7.96
6.51
6.80
7.30
8.29
8.82
9.01
7.83
7.79
7.75
7.76
7.75
7.75
8.34
8.28
8.43
8.16
8.60
8.50
7.70
10.0
10.4
10.7
9.8
9.8
7.9
13.9
13.6
12.8
12.5
12.4
12.5
16.1
16.7
19.0
19.1
19.2
14.1
14.1
14.0
14.1
13.9
14.5
12.8
12.9
12.5
12.5
12.7
13.0
5.0
12.8
3.0
14.2
3.3
14.9
3.6
31.85
17.75
20.18
25.82
19.46
20.53
18.14
17.34
26.49
27.02
36.73
39.25
34.17
28.60
25.36
26.44
23.21
157.35
94.05
74.20
13.85
3.95
2.51
36.49
40.88
36.97
39.08
32.09
40.58
17.32
15.40
11.86
23.39
15.27
10.09
38.52
16.77
10.15
11.55
13.59
10.41
10.46
21.94
21.80
23.66
26.01
29.91
35.05
25.87
22.87
21.42
22.34
21.52
27.18
18.82
23.78
24.21
18.62
16.19
26.65
27.80
23.44
25.22
20.34
25.73
33.37
26.39
27.20
31.76
48.41
26.48
22.41
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston & Russo 1983
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
West 1985
113
-------
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Oncorhynchus mykiss
Salmo trutta
Salmo trutta
Salmo trutta
Salvelinus fontinalis
Salvelinus fontinalis
Prosopium Williamson!
Prosopium Williamson!
Prosopium Williamson!
Notemigonus crysoleucas
Notropis lutrensis
Notropis lutrensis
Notropis spilopterus
Notropis spilopterus
Notropis spilopterus
Notropis whipplei
Campostoma anomalum
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
0.61
0.59
0.43
1.04
0.701
0.677
0.597
1.05
0.962
0.473
0.358
0.143
0.72
2.83
3.16
1.20
1.62
1.35
1.25
1.72
1.59
1.50
1.10
0.754
0.908
2.73
2.59
0.832
2.33
2.17
1.61
1.27
0.775
1.51
1.85
1.73
7.70
7.90
7.90
8.30
7.86
7.82
7.85
7.83
7.86
7.84
7.80
7.68
7.50
8.30
9.10
7.95
8.15
7.90
7.90
7.80
8.05
7.91
7.89
7.64
7.68
8.03
8.06
7.67
8.05
8.05
7.94
7.76
7.66
7.87
7.83
7.91
9.8
11.3
16.2
18.7
13.8
14.2
13.2
13.8
13.6
12.4
12.3
12.1
24.5
24.0
24.0
26.5
26.5
25.7
25.7
25.7
14.0
16.3
13.1
13.6
13.5
22.1
22.0
13.9
13.0
13.6
19.1
19.0
13.4
15.8
22.0
18.9
55.15
30.15
15.23
12.75
32.46
33.30
29.58
52.03
45.21
25.47
21.27
11.33
34.73
24.37
6.50
18.52
16.27
24.52
22.71
38.97
47.29
51.55
50.16
58.40
64.69
47.60
42.58
58.84
74.65
66.48
42.26
50.28
58.23
58.91
50.58
49.26
32.09
25.01
12.63
22.72
25.02
23.89
22.39
38.00
34.86
18.94
14.72
6.38
14.67
43.43
47.99
16.85
21.67
20.34
18.83
26.97
51.97
43.55
40.85
30.74
36.40
50.35
47.69
32.55
82.04
73.06
37.75
32.44
31.68
46.25
36.94
41.62
West 1985
West 1985
West 1985
West 1985
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Hazel etal. 1979
Hazel etal. 1979
Rosage et al. 1 979
Rosage et al. 1 979
Swigert & Spacie 1 983
Swigert & Spacie 1 983
Swigert & Spacie 1 983
DeGraeve et al. 1980
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
114
-------
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Catostomus commersoni
Catostomus commersoni
Catostomus commersoni
Catostomus commersoni
Catostomus commersoni
Catostomus commersoni
Catostomus commersoni
1.22
1.31
2.16
2.73
3.44
2.04
1.23
1.10
1.73
2.03
1.09
0.796
1.34
0.240
0.452
1.08
0.793
1.68
1.47
0.73
1.24
0.80
1.65
1.75
1.87
2.41
1.83
1.97
2.55
1.40
1.35
0.79
0.76
1.87
1.73
2.22
111
111
8.04
8.08
8.16
7.88
7.68
7.63
7.76
7.84
7.76
7.74
7.91
6.51
7.01
7.82
7.83
8.51
9.03
8.46
8.02
8.26
8.16
7.78
7.80
7.90
8.10
8.00
8.10
8.16
8.14
7.80
7.80
8.10
8.20
8.20
14.3
14.1
22.2
21.4
21.4
21.7
12.9
13.2
12.9
21.7
13.1
12.8
15.9
13.0
13.8
12.0
11.8
13.5
13.2
4.1
23.9
4.6
25.2
25.9
25.6
3.4
12.1
17.1
26.1
15.0
15.4
22.5
3.6
11.3
12.6
15.3
66.71
72.71
36.59
44.76
47.39
50.95
91.71
89.85
107.53
55.43
66.73
52.17
47.43
259.96
145.89
62.72
45.71
18.88
5.94
18.54
19.55
30.57
17.65
40.89
42.65
229.72
56.07
52.22
29.23
30.28
29.65
22.30
89.57
60.86
40.85
43.01
43.80
47.74
39.45
52.10
64.35
40.74
51.60
46.53
69.38
41.22
43.05
32.51
40.07
44.91
34.27
45.00
33.39
50.50
39.51
45.05
20.29
50.41
23.96
27.32
29.53
190.54
67.81
52.19
35.35
41.11
38.73
15.44
62.00
73.60
59.94
63.10
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1983
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
Thurston et al. 1981c
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
Swigert & Spacie 1 983
Swigert & Spacie 1 983
West 1985
West 1985
West 1985
West 1985
Reinbold & Pescitelli 1982c
Reinbold & Pescitelli 1982c
Swigert & Spacie 1 983
West 1985
West 1985
West 1985
West 1985
115
-------
Catostomus platyrhynchus
Catostomus platyrhynchus
Catostomus platyrhynchus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Ictalurus punctatus
Gambusia affinis
Gambusia affinis
Gambusia affinis
Gambusia affinis
Poecilia reticulata
Poecilia reticulata
Poecilia reticulata
Morone americana
Morone americana
Lepomis cyanellus
Lepomis cyanellus
Lepomis cyanellus
Lepomis cyanellus
Lepomis cyanellus
Lepomis cyanellus
Lepomis gibbosus
Lepomis gibbosus
0.819
0.708
0.668
2.4
2.9
3.8
1.95
2.1
4.2
1.76
1.75
1.45
0.50
0.98
1.91
1.29
2.26
2.6
2.4
3.2
2.4
1.47
1.59
1.45
0.15
0.52
0.61
1.08
0.594
1.29
1.64
2.11
0.14
0.78
7.67
7.73
7.69
8.70
8.70
8.70
8.40
8.09
8.08
7.98
7.94
7.80
8.00
8.10
8.10
7.80
8.00
8.00
8.20
7.75
8.50
7.22
7.45
7.45
6.00
8.00
7.84
8.28
6.61
7.20
7.72
8.69
111
111
12.0
11.7
13.2
22.0
26.0
30.0
28.0
22.0
28.0
23.8
23.8
25.7
3.5
14.6
17.0
19.6
26.0
24.0
19.5
19.0
23.0
25.0
25.0
25.0
16.0
16.0
12.3
26.2
22.4
22.4
22.4
22.4
12.0
14.5
66.91
51.62
47.59
10.56
10.19
10.88
10.71
32.33
44.44
30.49
33.10
32.85
37.64
24.94
40.83
44.71
32.34
42.53
34.54
129.59
14.64
129.40
82.95
75.65
418.44
14.93
33.09
8.43
254.49
142.85
55.79
9.24
9.11
42.02
37.02
31.62
27.23
40.26
38.85
41.47
23.19
38.36
51.72
29.35
29.57
22.74
37.61
30.16
49.38
30.95
32.32
42.51
50.68
82.17
38.41
37.66
32.56
29.69
63.94
14.92
24.61
14.45
45.86
40.64
33.59
34.60
5.98
27.59
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Thurston & Meyn 1 984
Colt & Tchobanoglous 1976
Colt & Tchobanoglous 1976
Colt & Tchobanoglous 1976
Colt & Tchobanoglous 1978
Roseboom & Richey 1 977
Roseboom & Richey 1 977
ReinboldS Pescitelli 1982b
ReinboldS Pescitelli 1982b
Swigert & Spacie 1 983
West 1985
West 1985
West 1985
West 1985
West 1985
Wallenetal. 1957
Wallenetal. 1957
Wallenetal. 1957
Wallenetal. 1957
RubinS Elmaraghy 1976,
1977
Rubin & Elmaraghy 1976,
1977
RubinS Elmaraghy 1976,
1977
Stevenson 1 977
Stevenson 1 977
Jude 1973
Reinbold & Pescitelli 1982a
McCormick et al. 1984
McCormick et al. 1984
McCormick et al. 1984
McCormick et al. 1984
Jude 1973
Thurston 1 981
116
-------
Lepomis gibbosus
Lepomis gibbosus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Lepomis macrochirus
Micropterus dolomieu
Micropterus dolomieu
Micropterus dolomieu
Micropterus dolomieu
Micropterus salmoides
Micropterus salmoides
Etheostoma spectabile
Etheostoma spectabile
Stizostedion vitreum
Stizostedion vitreum
Stizostedion vitreum
Stizostedion vitreum
Cottus bairdi
0.86
0.61
0.89
2.97
2.57
0.55
0.68
1.1
1.8
0.50
1.98
0.26
1.35
0.94
1.35
1.75
1.76
0.694
1.01
1.20
1.78
1.0
1.7
0.90
1.07
0.85
0.52
1.10
0.51
1.39
111
7.71
8.11
8.24
8.75
8.07
8.00
7.93
8.20
8.40
8.12
8.16
8.09
7.60
7.80
7.60
7.80
6.53
7.16
7.74
8.71
7.96
8.04
8.40
8.10
8.08
7.90
7.70
8.30
8.02
14.0
15.7
18.5
18.5
18.5
22.0
22.0
22.0
28.0
4.0
25.0
4.5
24.8
21.7
24.2
26.5
26.6
22.3
22.3
22.3
22.3
22.0
28.0
21.0
22.0
18.2
3.7
11.1
19.0
12.4
48.09
34.43
16.73
42.01
12.70
8.85
12.75
24.08
14.81
14.64
23.37
12.55
17.22
44.03
33.88
58.69
37.52
359.93
123.43
39.30
7.56
20.48
19.59
7.65
16.12
17.43
48.37
89.93
6.12
49.83
31.58
20.38
20.62
66.62
52.95
10.10
12.74
21.11
21.72
31.68
29.37
17.04
20.43
21.72
23.45
28.95
25.97
62.67
33.60
24.49
29.33
18.99
21.12
16.55
19.49
20.29
40.12
52.33
10.91
51.73
Thurston 1 981
Thurston 1 981
Emery & Welch 1969
Emery & Welch 1969
Emery & Welch 1969
Roseboom & Richey 1 977
Roseboom & Richey 1 977
Roseboom & Richey 1 977
Roseboom & Richey 1 977
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
ReinboldS, Pescitelli 1982b
Smith etal. 1983
Swigert & Spacie 1 983
Swigert & Spacie 1 983
Swigert & Spacie 1 983
Broderius et al. 1985
Broderius et al. 1985
Broderius et al. 1985
Broderius et al. 1985
Roseboom & Richey 1 977
Roseboom & Richey 1 977
Hazel etal. 1979
Hazel etal. 1979
Reinbold & Pescitelli 1982a
West 1985
West 1985
West 1985
Thurston S, Russo 1981
The species and tests are in the same order as in Table 1 in the 1984/1985 ammonia criteria
document. The scientific names of various salmonids have been updated. Two values for the
rainbow trout by Calamari et al. (1977,1981) were deleted because they were "greater than"
values; this had no effect on the FAV because the SMAV for rainbow trout was lowered to protect
large rainbow trout (see Table 4 in this 1999 Update). A few values for pH and temperature were
corrected and ranges were replaced with point estimates to facilitate conversion of acute values
from un-ionized ammonia at the test pH to total ammonia nitrogen at pH=8.
117
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Appendix 5. Histopathological Effects
Fewer results of the effects of chronic exposure of aquatic life to ammonia are available
than results of the effects of acute exposures. The available data indicate that
ammonia can have adverse effects on aquatic life at relatively low concentrations,
approaching 0.001 to 0.006 mg NH3-N/L. These reported adverse effects include
quantitative data showing that decreased survival, growth, and reproduction are
correlated to increasing concentrations of ammonia. These more conventional
measures of chronic toxicity are generally regarded as a suitable basis for projecting
the potential chronic toxic effects of pollutants, including ammonia, to aquatic life
populations and communities.
In addition to the reported chronic toxic effects of ammonia to aquatic life based on
these more conventional measures, the literature contains some information
concerning the effects that chronic exposure to low levels of ammonia can have on the
structure and function of select tissues and organs. These include reduced swimming
stamina and performance, increased respiratory distress, hormonal dysfunction, and
damage to gill, kidney, brain, and liver tissues. Some investigators have reported other
pathological changes in the test animals' physiology, histochemistry, and biochemistry.
None of these reported abnormalities in test organisms have been quantitatively
correlated with the ammonia exposure or with effects on the survival, growth, or
reproduction of the test organisms; potential adverse effects on populations and
communities are unavailable.
Salmonid species subjected to un-ionized ammonia concentrations ranging from 0.002
mg NH3-N/L at pH=6.4 to 0.06 mg NH3-N/L at pH=7.7 on a chronic exposure basis have
demonstrated significant effects on growth. Rice and Bailey (1980) observed growth
effects on pink salmon embryos and fry when un-ionized ammonia exceeded 0.002 to
0.003 mg NH3-N/L at pH=6.4. Burkhalter and Kaya (1977) observed that un-ionized
ammonia concentrations somewhat less than 0.05 mg NH3-N/L at pH=7.5 inhibited
growth rates of rainbow trout embryos and fry. Samylin (1969), in tests with Atlantic
salmon embryos and fry, reported effects on growth rates when un-ionized ammonia
exceeded 0.06 mg NH3-N/L at pH=7.1. The calculated "no apparent effect"
concentrations for these tests are 0.002 mg NH3-N/L at pH=6.4 for pink salmon, 0.008
mg NH3-N/L at pH=7.1 for the Atlantic salmon, and less than 0.05 mg NH3-N/L at
pH=7.5 for the rainbow trout. Non-salmonid fish species have exhibited similar effects,
with the calculated "no apparent growth effect" concentrations ranging from 0.03 mg
NH3-N/L at pH=6.6 to 0.05 mg NH3-N/L at pH=8.68. Reported growth effect
concentrations were 0.11 mg NH3-N/L at pH=7.78forthe bluegill (Smith et al. 1984),
0.32 mg NH3-N/L at pH=7.95 for the channel catfish (Reinbold and Pescitelli 1982a),
and 0.40 mg NH3-N/L at pH=7.9 for the green sunfish (McCormick et al. 1984).
Broderius et al. (1985), in tests with smallmouth bass, observed that the growth effects
of un-ionized ammonia were not constant with pH. The growth effect concentrations
ranged from 0.05 mg NH3-N/L at pH=6.6 to 0.71 mg NH3-N/L at pH=8.68. Thurston et
al. (1986) reported the results of life-cycle tests with the fathead minnow. The tested
118
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un-ionized ammonia concentrations ranged from 0.07 to 0.96 mg NH3-N/L at pH=8.0.
No effects on growth or survival of parental fish were reported at 0.44 mg NH3-N/L, or
on embryo viability or production up to 0.37 mg NH3-N/L; adverse effects were reported
for all of these endpoints at 0.91 mg NH3-N/L. First filial generation animals did not
demonstrate any adverse effects on growth or survival at 0.36 mg NH3-N/L, the highest
tested concentration. Embryo hatching success was adversely affected at 0.37 mg
NH3-N/L but not at 0.19 mg NH3-N/L. Parental fish and first filial generation fish
exhibited a high incidence of brain lesions at an un-ionized ammonia concentration of
0.21 mg NH3-N/L, but not at 0.11 mg NH3-N/L
Histopathological effects of chronic exposure of rainbow trout to un-ionized ammonia
are evident within the range of un-ionized concentrations producing effects on growth.
Calamari et al. (1977,1981) observed alterations of the epidermis of newly hatched
rainbow trout fry exposed to un-ionized ammonia concentrations of 0.02 mg NH3-N/L
and greater at pH=7.4 for 21 to 24 days. Concentrations of 0.06 mg NH3-N/L and
greater at pH=7.4 produced pathological alterations of kidney tissues of newly hatched
rainbow trout fry. Increases in the severity of these pathological states corresponded
to increasing un-ionized ammonia concentrations; fifty percent mortality was reported
with animals exposed to concentrations of 0.06 mg NH3-N/L and greater at pH=7.4 for
72 days (Calamari et al. 1977,1981).
Thurston et al. (1984b) exposed rainbow trout to five concentrations of un-ionized
ammonia ranging from 0.008 to 0.06 mg NH3-N/L at pH=7.7. The parental (P) fish were
exposed for eleven months, the first filial generation (F.,) for 48 months, and the second
filial generation (F2) for five months. Animals from the parental, first filial, and second
filial generations were examined for chronic effects of un-ionized ammonia. Data
collected during the tests included mortality, reproductive success, and growth.
Histological examinations were performed on select tissues from fish of all three
generations.
No statistically significant difference in survival, growth, or reproduction was observed
at any of the tested concentrations. Blood from the parental fish exposed to
concentrations of 0.05 mg NH3-N/L and greater showed reduced hematocrits and, to a
lesser extent, reduced hemoglobin content. The first filial generation (F.,) did not show
any significant alteration in hematocrits or hemoglobin, although there was a strong
correlation between blood ammonia values and ambient ammonia concentrations.
Histological examinations of spleen, heart, gill, liver, and kidney tissues were
performed on animals from all three generations and correlated to test concentrations.
Histological alterations of gill and kidney tissues were remarkable and showed a
positive correlation with un-ionized ammonia concentrations; histopathological
alterations increased in severity with increasing ammonia concentrations. Gill lamellae
obtained from parental fish exposed to un-ionized ammonia concentrations ranging
from 0.02 mg NH3-N/L to 0.05 mg NH3-N/L for four months, and 0.05 mg NH3-N/L and
0.06 mg NH3-N/L for seven and eleven months, showed mild to moderate fusion,
aneurysms, and separation of the epithelia from the underlying basement membrane.
119
-------
Test animals that had been exposed for seven months at un-ionized ammonia
concentrations of 0.05 mg NH3-N/L and subsequently allowed to 'recover' in an
ammonia-free environment for the remaining four months, did not show any evidence of
gill tissue damage, suggesting that the animals might have recovered.
The gill tissues offish from the first filial generation exposed to concentrations of 0.03
mg NH3-N/L and greater evidenced mild to severe tissue injury. The degree of injury
exhibited a positive correlation with the un-ionized ammonia concentrations.
Symptoms included hypertrophy of the gill lamellae, with accompanying basal
hyperplasia, separation of epithelia from the underlying basement membranes,
necrosis, aneurysms, and mild to moderate fusion of gill lamellae. This suite of
symptoms is analogous to obstructive bronchopulmonary disease, e.g., emphysema, in
humans and has been reported to affect swimming performance and stamina in trout
(Smith and Piper 1985). Pathologic conditions were most apparent in both the parental
and P., fish when un-ionized ammonia reached and exceeded 0.03 mg NH3-N/L at
pH=7.7. No effects were reported on survival, growth, or reproduction at the highest
tested concentration of 0.06 mg NH3-N/L.
Second filial generation rainbow trout exposed to un-ionized ammonia concentrations
of 0.02 mg NH3-N/L and greater exhibited histological alterations similar to those of the
first filial generation. In addition to the histopathological alterations, the second filial
generation also became infected with a protozoan. It is not known whether the
protozoan infection was related to an increased susceptibility associated with the
ammonia exposure. These alterations are generally viewed as pathological and
strongly indicative of organ dysfunction. Survival and growth of the second filial
generation were unaffected at the highest tested ammonia concentration of 0.06 mg
NH3-N/L.
In addition to the recovery noted by Thurston et al. (1984b), other investigators have
reported recovery and compensation. Smith and Piper (1975) reported recovery of
rainbow trout when in water to which ammonia was not added. Burrows (1964)
observed recovery of Chinook salmon in uncontaminated water at 14°C, but not at 6°C.
Schulze-Wiehenbrauck (1976) found that growth of rainbow trout juveniles was
reduced during two-week exposures, but the decrease was completely compensated
for during the next three or four weeks. Burkhalter and Kaya (1977) reported
compensation for reduced growth at the lowest tested concentration.
Endpoint indices of abnormalities such as reduced growth, impaired reproduction,
reduced survival, and gross anatomical deformities are clinical expressions of altered
structure and function that originate at the cellular level. Any lesion observed in the
test organism is cause for concern and such lesions often provide useful insight into
the potential adverse clinical and subclinical effects of such toxicants as ammonia. For
purposes of protecting human health or welfare these subclinical manifestations often
serve useful in establishing 'safe' exposure conditions for certain sensitive individuals
within a population.
120
-------
With fish and other aquatic organisms the significance of the adverse effect can be
used in the derivation of criteria only after demonstration of adverse effects at the
population level, such as reduced survival, growth, or reproduction. Many of the data
indicate that the concentrations of ammonia that have adverse effects on cells and
tissues do not correspondingly cause adverse effects on survival, growth, or
reproduction. No data are available that quantitatively and systematically link the
effects that ammonia is reported to have on fish tissues with effects at the population
level. This is not to say that the investigators who reported both tissue effects and
population effects within the same research did not correlate the observed tissue
lesions and cellular changes with effects on survival, growth, or reproduction, and
ammonia concentrations. Many did, but they did not attempt to relate their
observations to ammonia concentrations that would be safe for populations of fish
under field conditions nor did they attempt to quantify (e.g., increase in respiratory
diffusion distance associated with gill hyperplasia) the tissue damage and cellular
changes (Lloyd 1980; Malins 1982). Additionally, for the purpose of deriving ambient
water quality criteria, ammonia-induced lesions and cellular changes must be
quantified and positively correlated with increasing exposures to ammonia.
In summary, the following have been reported:
1. Fish recover from some histopathological effects when placed in water that does
not contain added ammonia.
2. Some histopathological effects are temporary during continuous exposure of fish to
ammonia.
3. Some histopathological effects have occurred at concentrations of ammonia that
did not adversely affect survival, growth, or reproduction during the same
exposures.
Because of the lack of a clear connection between histopathological effects and effects
on populations, histopathological endpoints are not used in the derivation of the new
criterion, but the possibility of a connection should be the subject of further research.
121
-------
Appendix 6. Results of Regression Analyses of Chronic Data
The following pages contain figures and other information related to the regression
analyses that were performed to calculate chronic EC20s and LC20s. The regression
parameters are shown in the table below. In the figures that follow, circles denote
measured responses and confidence limits (if available), solid lines denote estimated
regression lines, and dotted lines denote 95% confidence limits on the regression lines.
Squares with solid thick lines denote estimated EC20s and 95% confidence limits.
Study
Fingernail Clam, Anderson et al.
Fingernail Clam, Sparks and Sandusky
Ceriodaphnia acanthina, Mount et al.
Ceriodaphnia dubia, Willingham
Ceriodaphnia dubia, Nimmo et al.
Daphnia magna, Gersich et al.
Daphnia magna, Reinbold and Pescitelli
Hyalella azteca, Borgmann
Fathead minnow, Thurston et al.
Fathead minnow, Swigert and Spacie
Fathead minnow, Mayes et al.
Channel catfish, Swigert and Spacie
Channel catfish, Reinbold and Pescitelli
Green sunfish, Reinbold and Pescitelli
Green sunfish, McCormick et al.
Bluegill, Smith etal.
Smallmouth bass, Broderius et al
Smallmouth bass, Broderius et al
Smallmouth bass, Broderius et al
Smallmouth bass, Broderius et al
Control
Value
84.6
92.8
12.6
23.0
12.7
67.8
24.3
48.0
49.3
2.81
70.5
7.29
61.2
81.7
2.84
2.10
168.
161.
123.
121.
EC50
8.00
2.69
49.8
7.53
25.3
15.3
27.6
1.41
4.56
13.1
7.25
14.1
45.8
6.57
8.78
2.56
22.5
26.6
12.4
2.29
Slope
4.35
1.78
13.3
5.32
2.72
1.89
5.77
2.90
1.65
1.11
3.98
6.41
1.06
11.8
3.09
4.27
1.63
1.23
3.30
3.47
EC20
4.54
1.23
41.5
5.80
15.2
7.37
21.7
0.88
1.97
3.73
5.12
11.5
12.4
5.84
5.61
1.85
9.61
8.62
8.18
1.54
EC10
3.46
0.78
37.3
4.98
11.3
4.81
18.9
0.66
1.20
1.79
4.18
10.1
5.77
5.45
4.31
1.53
5.84
4.46
6.40
1.22
122
-------
FINGERNAIL CLAM, 42-DAY JUV, ANDERSON ET AL. 1978
03
C/5
100
80
60
c
O 40
0)
Q_
20
EC20 = 5.82 mg N/L (4.54-7.46)
T = 23.5°C
pH = 8.15
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
FINGERNAIL CLAM, 42-DAY JUV, SPARKS AND SANDUSKY 1981
03
C/5
"c
CD
CD
Q.
100 r
80
60
40
20
EC20 = 1.23 mg N/L (0.86-1.76)
T = 21.8°C
pH = 7.80
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
10 20
123
-------
CERIODAPHNIA ACANTHINA, LIFE CYCLE, MOUNT 1982
15
E
to
'c
03
O)
6
"to
CD
co
c
_
O
i_
CD
Q.
O)
C
12
EC20 = 44.9 mg N/L (41.5-48.6)
T = 24.5°C
pH = 7.15
0.5
2 5 10 20
Total Ammonia (mg N/L)
100
CERIODAPHNIA DUBIA, LIFE CYCLE, WILLINGHAM 1987
30
C 25
CO
s>
220
to
CD
15
O
i_
CD
Q.
O)
C
10
fll—L
EC20 = 5.80 mg N/L (4.12-8.15)
T = 26.0°C
pH = 8.57
0.1 0.2 0.5 1 2 5 10
Total Ammonia (mg N/L)
20
50
124
-------
CERIODAPHNIA DUBIA, LIFE CYCLE, NIMMO ET AL 1989
E
CO
'c
03
oo
CD
03
C
i_
CD
Q.
O)
§
O
15
12
EC20 = 15.2 mgN/L (9.3-24.8)
T = 25°C
pH = 7.8
0.5
2 5 10 20
Total Ammonia (mg N/L)
50
100
DAPHNIA MAGNA, LIFE CYCLE, GERSICH ET AL. 1985
80
60
E
CO
'c
03
co
CD
g 40
_
O
_
CD
Q.
O)
C
20
EC20 = 7.37 mg N/L (4.13-13.15)
T= 19.8°C
pH = 8.45
0.1 0.2 0.5 1 2 5 10
Total Ammonia (mg N/L)
20
50
125
-------
DAPHNIA MAGNA, LIFE CYCLE, REINBOLD AND PESCITELLI 1982a
30
I 25
03
O)
° 90
^ 20
to
CD
15
O
1_
CD
Q.
O)
c
10
EC20 = 21.7 mg N/L (12.1-39.2)
T = 20.1°C
pH = 7.92
0.2 0.5 1 2 5 10
Total Ammonia (mg N/L)
20
50
HYALELLA AZTECA, LIFE CYCLE, BORGMANN 1994
EC20 = 0.88 mg N/L (0.58-1.32)
T = 25°C
pH = 8.04
0.5 1 2 5
Total Ammonia (mg N/L)
10
20
126
-------
FATHEAD MINNOW, LIFE CYCLE, THURSTON ET AL. 1986
60 h
03
CD
^
CD
Q.
50
40
30
20
10
EC20 = 1.97 mgN/L (0.99-3.91)
T = 24.2°C
pH = 8.00
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
10
20
FATHEAD MINNOW, 30-DAY ELS, SWIGERT AND SPACIE 1983
3.0
—. 2.5
O)
to
03 2.0
T3
O
CO
03
1.5
to
to
03
E 1.0
o
m
0.5
EC20 = 3.73 mg N/L (1.55-8.98)
T = 25.1°C
pH = 7.82
0.0
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
10
20
127
-------
FATHEAD MINNOW, 28-DAY ELS, MAYES ET AL. 1986
03
C/5
100 r
80
60
c
O 40
Q_
20
EC20 = 5.12 mg N/L (4.27-6.14)
T = 24.8°C
pH = 8.00
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
100 r
80
"03
'> 60
co
O 40
CD
Q.
20
O1—>-
0.1
EC20 = 12.2 mg N/L (8.2-18.1)
T = 6.0°C
pH = 7.83
J I L
J L
0.2 0.5 1
5 10 20 50 100
Total Ammonia (mg N/L)
128
-------
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
CO
l_
C/5
-t->
CD
CD
Q.
100 r
80
60
40
20
EC20 = 18.0 mgN/L (5.4-60)
T= 10.0°C
pH = 7.73
0.1 0.2 0.5 1 2 5 10 20
Total Ammonia (mg N/L)
i-
i
50 100
CO
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
100 r « • - • <
80
60
C/5
"c
O 40
CD
Q.
20
EC20 = 39 mg N/L (29-52)
T = 25.4°C
pH = 7.35
0.1 0.2 0.5 1 2 5 10 20
Total Ammonia (mg N/L)
50 100
129
-------
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
03
C/5
100 r
80
60
c
O 40
Q_
20
O'-L-
0.1
EC20 = 35 mg N/L (17-72)
T = 30.2°C
pH = 7.19
0.2 0.5 1 2 5 10 20
Total Ammonia (mg N/L)
50 100
CHANNEL CATFISH, 30-DAY ELS, SWIGERT AND SPACIE 1983
10 r
O)
03 _
-a 6
o
CO
"03
03
E
o
m
EC20 = 11.5mg N/L (9.7-13.6)
T = 26.9°C
pH = 7.76
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
10
20
130
-------
CHANNEL CATFISH, 30-DAY ELS, REINBOLD AND PESCITELLI 1982a
70 r
O) 60
CO
T3
3 50
"co
g>
40
30
EC20 = 12.2 mgN/L (4.3-28.9)
T = 25.8°C
pH = 7.80
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
10
20
GREEN SUNFISH, 30-DAY ELS, REINBOLD AND PESCITELLI 1982a
100 r
80
"co
'> 60
C/5
O 40
CD
Q.
20
EC20 = 5.84 mg N/L (5.07-6.72)
T = 25.4°C
pH = 8.16
0.1 0.2 0.5 1 2 5
Total Ammonia (mg N/L)
20
131
-------
GREEN SUNFISH, 30-DAY ELS, MCCORMICK ET AL. 1984
O)
co
CO
03
E
o
m
EC20 = 5.61 mg N/L (2.84-11.06)
T = 22.0C
pH = 7.9
0.2 0.5 1 2 5
Total Ammonia (mg N/L)
20
BLUEGILL, 30-DAY ELS, SMITH ET AL. 1984
2.5 r
2.0
5
co
CO
03
1.5
0.5
0.0
EC20 = 1.85 mg N/L (1.42-2.42)
T = 22.5°C
pH = 7.76
0.2 0.5 1 2 5 10 20
Total Ammonia (mg N/L)
132
-------
SMALLMOUTH BASS, 32-DAY ELS, BRODERIUS ET AL. 1985
200 r
D) 150
CD
Q
co
100
co
(§ 50
EC20 = 9.61 mg N/L (6.59-14.02)
T = 22.3°C
pH = 6.60
0.5
2 5 10
Total Ammonia (mg N/L)
20
50
SMALLMOUTH BASS, 32-DAY ELS, BRODERIUS ET AL. 1985
200 r
D) 150
co
Q
co
100
co
o
50
EC20 = 8.62 mg N/L (5.57-13.36)
T = 22.3°C
pH = 7.25
0.5
2 5 10
Total Ammonia (mg N/L)
20
50
133
-------
SMALLMOUTH BASS, 32-DAY ELS, BRODERIUS ET AL. 1985
150
D)
CD
Q
CM
CO
"co
co
E
o
in
100
50
EC20 = 8.18 mg N/L (5.89-11.37)
T = 22.3°C
pH = 7.83
0.2
0.5 1 2 5
Total Ammonia (mg N/L)
10
20
150
SMALLMOUTH BASS, 32-DAY ELS, BRODERIUS ET AL.
D)
co
Q
CM
CO
-i-»
co
co
o
in
100
50
EC20 = 1.54 mg N/L (1.25-1.89)
T = 22.3°C
pH = 8.68
o
0.05
0.1 0.2 0.5 1 2
Total Ammonia (mg N/L)
10
134
-------
Appendix 7. Acute-Chronic Ratios
Although the CCC was calculated directly from Chronic Values using the fifth percentile
procedure (U.S. EPA 1985b), it is of interest to consider how this compares with the
use of Acute-Chronic Ratios (ACRs). Therefore, ACRs were determined for all of the
EC20s in Table 5 that are used in the derivation of a GMCV and for which comparable
acute values were found. (Sufficient ACRs are available for freshwater species that
ACRs determined with saltwater species were not considered.) Because the acute
toxicity of total ammonia is related to pH differently from its chronic toxicity, all relevant
acute and chronic values were adjusted to pH=8 and are expressed in terms of mg N/L,
where N is total ammonia nitrogen. The resulting ACRs are given in Table 7, along
with the resulting Genus Mean Acute-Chronic Ratios (GMACRs).
When ACRs are used, it is hoped that if the acute and chronic tests are conducted with
the same test species in the same water, any biological or chemical factor that affects
the result of one of the tests will have a proportional effect on the result of the other test
so that the ACR is more constant than the result of either individual test. In addition, it
is hoped that the ACRs within a genus agree well. The ACRs within the genera
Ceriodaphnia and Daphnia agree well (Table 7).
The available ACRs at pH=8 for the fathead minnow range from 6.5 to 20.7, but the
range can probably be explained because of the different kinds of chronic tests on
which they are based. The ACR of 20.7 was based on the life-cycle test of Thurston et
al. (1986) whereas the early life-stage tests of Swigert and Spacie (1983) and Mayes et
al. (1986) gave ACRs of 6.5 and 9.7. The range of ACRs for the early life-stage tests is
small, and it is not surprising that a life-cycle test gave a higher ACR than the early life-
stage test. The range of the nine 96-hr LCSOs from three laboratories was only 27.2 to
51.5 mg N/L when adjusted to pH=8.
Table 8 gives the GMACRs beside the ranked GMAVs to demonstrate whether there is
a trend, because ACRs for some chemicals are higher for resistant species than for
sensitive species (U.S. EPA 1985b). No trend is obvious and the range of the
GMACRs is 1.9 to 10.9.
A major problem with use of the ACR procedure for calculating a CCC for ammonia is
that ACRs are not available for M. transversum and H. azteca, which are very sensitive
in chronic tests; the data in the 1984/1985 ammonia criteria document indicate that M.
transversum is not very sensitive in acute tests, which implies a large ACR. In these
circumstances, direct calculation of the CCC using the fifth percentile calculation
procedure is certainly much more appropriate than calculation using the ACR
procedure. In addition, the CCC obtained using the fifth percentile procedure agrees
well with the available chronic data.
135
-------
Table 7. Genus Mean Acute-Chronic Ratios
Species Chronic Results Acute Results3 Adjusted to pH=8
Mtrsncworci im
. LI a! loVCI oUI 1 1
C. acanthina
C. dubia
D. magna
Ho7tor*3
. aZ.LCUa
P. promelas
C. commersoni
1. punctatus
L. cyanellus
Refb
1
I
9
£-
3
4
5
6
7
9
10
11
7
10
7
12
Temp
9-2 K
^o.o
91 ft
^ i .0
24.5
26.0
25.0
19.8
20.1
OK n
^o.u
24.2
25.1
24.8
18.6
26.9
25.4
22.0
PH
81^
. I O
7 ftO
/ .ou
7.15
8.57
7.8
8.45
7.92
7 Q4
/ . yt
8.0
7.82
8.0
8.32
7.76
8.16
7.9
EC20C
K QO
O.O^
1 9^
I .^O
44.9
5.80
15.2
7.37
21.7
<-1 KQ
^ 1 .OO
1.97
3.73
5.12
>2.9
11.5
5.84
5.61
Temp
24.0
26.0
25.0
20.0
19.7
22.1
22.0
19.1
19.0
22.0
18.9
25.9
25.6
22.0
15.0
15.4
25.7
26.2
22.4
PH
7.06
8.61
7.8
8.50
8.34
8.03
8.06
7.94
7.76
7.83
7.91
7.78
7.8
8.14
8.16
8.14
7.8
8.28
7.7
LC50C
105.
14.8
41.3
26.4
61.3
48.6e
42.6e
42. 3e
50.4e
50.6e
49.3e
41.0
42.8
25.2
30.3f
29.7f
32.8
8.6
57.
EC20
7 ^n
/ .ou
n Q4
u. yt
19.8
14.1
11.6
15.1
19.4
<1 4^
^ I .HO
1.97
2.92
5.12
>4.79
8.35
7.44
4.88
LC50
24.4
48.6
31.5
70.2
119.
51.5
47.8
37.7
32.2
36.7
41.5
27.2
29.4
33.1
41.4
39.0
22.6
14.8
32.8
ACRd
1.2
3.4
2.7
4.6
6.1
20.7
9.7
6.5
<8.4
2.7
2.0
6.7
GM/>
1.9
5.3
10.9
<8.4
2.7
7.6
L. macrochirus 13 22.5 7.76 1.85 21.7 7.6 44.2 1.35 21.4 15.9
136
-------
M. dolomieu 14 22.3 6.60 9.61 22.3 6.53 371. 3.57 59.3 16.6 7.4
22.3 7.25 8.62 22.3 7.16 117. 4.01 30.4 7.6
22.3 7.83 8.18 22.3 7.74 39.5 6.50 24.4 3.8
22.3 8.68 1.54 22.3 8.71 7.43 4.65 29.3 6.3
a If acute values were available at more than one pH, the acute value(s) at a pH close to the pH of the chronic value were used.
Dashes indicate that a comparable acute test was not found. When an acute test listed above was in Table 1 of the 1984/1985
ammonia criteria document (U.S. EPA 1985a), the values given in Table 1 for pH and temperature were used unless inspection
of the reference indicated that an incorrect value was in Table 1. If given in the reference, an LC50 based on total ammonia
was used, after conversion to total ammonia nitrogen if necessary. If a total ammonia LC50 was not given in the reference, an
LC50 based on un-ionized ammonia was used, after conversion to un-ionized ammonia nitrogen if necessary. Each LC50
based on un-ionized ammonia nitrogen was converted to total ammonia nitrogen in the table above, using the speciation
relationship derived by Emerson et al. (1978).
b (1) Anderson et al. 1978; (2) Sparks and Sandusky 1981; (3) Mount 1982; (4) Willingham 1987; (5) Nimmo et al. 1989; (6)
Gersich et al. 1985; (7) Reinbold and Pescitelli 1982a; (8) Borgmann 1994; (9) Thurston et al. 1986; (10) Swigert and Spacie
1983; (11) Mayesetal. 1986; (12) McCormick et al. 1984; (13) Smith etal. 1984; (14) Broderius et al. 1985.
c Expressed as total ammonia nitrogen (mg N/L). Three digits are retained in intermediate calculations to reduce roundoff error in
subsequent calculations.
d One ACR was calculated for each EC20 for which a comparable acute value was available; if more than one comparable acute
value was available, the geometric mean of the acute values was used.
e These are the results of the six acute tests given by Thurston et al. (1983) in their appendix that were conducted with fish that
were 0.1 to 1.0 g and whose test temperature was closest to the temperature of the chronic test.
f Reinbold and Pescitelli 1982b.
137
-------
Table 8. Ordered Genus Mean Acute-Chronic Ratios
RANK
34
33
32
31
30
29
28
27
26
25
24
23
22
21
20
19
18
17
16
15
14
13
12
11
10
9
8
7
6
5
4
3
2
1
GENUS
Philarctus
Orconectes
Asellus
Ephemerella
Callibaetis
Stenelmis
Crangonyx
Tubifex
Helisoma
Arcynopteryx
Physa
Cottus
Gambusia
Pimephales
Catostomus
Daphnia
Salvelinus
Musculium
Ictalurus
Simocephalus
Poecilia
Dendrocoelum
Morone
Campostoma
Micropterus
Stizostedion
Ceriodaphnia
Notropis
Salmo
Lepomis
Oncorhynchus
Etheostoma
Notemigonus
Prosopium
GMAV ADJUSTED TO pH=8
388.8
246.0
210.6
189.2
115.5
113.2
108.3
97.82
93.52
77.10
73.69
51.73
51.06
43.55
38.11
36.82
36.39
35.65
34.44
33.99
33.14
32.82
30.89
26.97
26.50
26.11
25.78
25.60
23.74
23.61
21.95
17.96
14.67
12.11
GMACR
10.9
<8.4
5.3
2.7
7.4
1.9
7.6
138
-------
Appendix 8. A Field Study Relevant to the CCC
Hermanutz et al. (1987) and Zischke and Arthur (1987) reported the effects of different
concentrations of ammonia on fishes and invertebrates in various tests at the
Monticello, MN, outdoor experimental stream facility. The study involved essentially
constant dosing of total ammonia into four parallel streams (three concentrations of
ammonia and a control treatment). The approximate average concentrations of total
ammonia nitrogen were:
0.08 mg N/L in the control stream
0.66 mg N/L in the low concentration stream
2.0 mg N/L in the medium concentration stream, and
7.1 mg N/L in the high concentration stream.
Although the streams were physically identical, the different concentrations of ammonia
caused chemical and microbiological differences among the streams. Higher ammonia
concentrations yielded lower pH, and, as a result of higher nitrifying bacterial activity,
higher nitrite and nitrate concentrations and lower concentrations of dissolved oxygen,
particularly in the lower reaches of the streams containing added ammonia. For
example, in the lower reaches of the high concentration stream, dissolved oxygen
regularly dropped to 2 mg/L at night during summer. Although these differences
between streams reflect real-world phenomena usually accompanying ammonia
enrichment, they confound interpreting some of the results in terms of the toxicity of
ammonia. Six of the thirteen tests with fishes, however, either did not use the lower
reaches of the streams or did not take place during the summer. For these tests the
confounding influences of nitrifier activity should not be of much concern.
The study began in June 1983 and ended in November 1984, but all of the tests with
the various taxa were of shorter durations. Macroinvertebrate tests lasted for two
months, whereas the durations of the fish tests were 28 to 237 days. During all of the
tests, the organisms were left to forage on naturally occurring flora and fauna, except
that the walleyes were fed fathead minnows.
As reported by Hermanutz et al. (1987), densities of individual macroinvertebrate taxa,
sampled approximately 1 to 2 months after the start of the dosing, differed somewhat
among the streams. Cladoceran and protozoan densities might have been inhibited by
elevated ammonia concentrations (or accompanying changes), rotifer densities might
have been somewhat stimulated, and copepod densities showed little effect. However,
concentration-effect patterns were generally inconsistent, and the results do not
support any overall conclusion of either stimulatory or inhibitory effects. Because
laboratory toxicity tests indicate that these types of macroinvertebrates are generally
substantially more resistant to ammonia than fishes, absence of effects might not be
viewed as unexpected.
Tests with fishes included two tests with the fathead minnow, one with the bluegill,
three with the channel catfish, two with the white sucker, two with the walleye, and
139
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three with the rainbow trout. Hermanutz et al. (1987) studied percent survival, fish
length, fish weight, and final fish biomass, and identified those treatments and variables
that were significantly different than the control stream for individual species. The
Technical Support Document for Water Quality-based Toxics Control, EPA (1991)
attempted a subjective summarization of these results, relative to the CCC defined in
U.S. EPA(1985a).
The fingernail clam data of Zischke and Arthur (1987) were also evaluated. These
investigators selected this species for study because it is an important component of
many freshwater communities and because it was reported to be highly sensitive to
ammonia (Anderson et al. 1978; Sparks and Sandusky 1981).
The intent of this new analysis is to provide a quantitative graphical portrayal of the
results of the thirteen tests with fishes and the two tests with the fingernail clam.
Recognizing that field and macrocosm data involve a substantial amount of variability,
this analysis is intended to determine whether any pattern emerged from the noise.
To integrate the results as much as possible, this analysis used biomass at the end of
each test with fish, which Hermanutz et al. (1987) determined from the number of
surviving individuals multiplied by the individual mean weight. For the fathead minnow,
this measure combines survival, growth, and reproduction. For the other tested fish
species, this measure combines survival and growth. Biomass was not available from
the data on the fingernail clam. In its place, the product of survival and mean organism
length was used.
Concentrations of ammonia were normalized to account for the dependence of
ammonia toxicity on pH. The exposure metric used was the concentration of ammonia
in the stream divided by the 1998 CCC, which is 2% lower than the 1999 CCC at 25°C
but 35% higher than the 1998 CCC at 20°C.
Because both 4-day and 30-day averaging periods are used in the criteria statement,
this analysis considered whether the maximum 4-day or the maximum 30-day average
was significantly different than the long-term average concentration. Although the
concentration of total ammonia varied little over the duration of the Monticello tests, the
pH, and therefore ammonia toxicity, varied somewhat over time, particularly in the
longer tests. In this case, the CCC varies over time, while the concentration of total
ammonia is more constant. The CCC calculated from the maximum 4-day mean pH
would be lower than the CCC calculated from the maximum 30-day mean pH. Both
would be lower than the CCC calculated from the long-term mean pH. Because the
original data books for these tests are no longer available, this analysis relied on data
published by Hermanutz et al. (1987) and Zischke and Arthur (1987), which precluded
any attempt to estimate the day-by-day exposure.
For tests of 28 to 90 days (that is, up to threefold greater than the 30-day averaging
period), the applicable CCC applied with a 30-day averaging period was calculated
from the mean pH for the test. For the longer tests within this range, use of the mean
pH probably causes a slight bias toward underestimating the excursion of the CCC.
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For tests of 91 to 237 days (more than threefold greater than the 30-day averaging
period), the applicable CCC applied with a 30-day averaging period was calculated
from the highest 30-day mean pH occurring during the test. For the high ammonia
stream, this mean pH was estimated directly from the published graph of pH-time
variability in this stream. For the other streams, which lacked published graphs on the
time course of pH variations, the maximum 30-day mean pH was estimated from the
test mean pH for the stream, coupled with the variation about the mean observed in the
high treatment stream. That is, the degree of pH variability was assumed to be the
same in all of the streams.
For the fish tests, the applicable CCC applied with a 4-day averaging period was
estimated from the maximum weekly mean pH, estimated from the published graphs, or
from the expected pH variability, in the manner described in the preceding paragraph.
For the fingernail clam tests, the maximum 4-day mean pH was taken to be the
maximum weekly mean pH published by Zischke and Arthur (1987) for their tests,
which is likely to be lower than the actual maximum 4-day mean pH.
Table 9 presents the fish data from Hermanutz et al. (1987) and the fingernail clam
data from Zischke and Arthur (1987). The results of the analysis are presented in
Figure 15, which show the biological effect, relative to the control treatment, on the
vertical axis, and the exposure concentration, relative to the 1998 CCC of 1.27 mg N/L,
on the horizontal axis.
Uncertainties exist in the vertical and horizontal locations of points in Figure 15.
Biological measurements on side-by-side macrocosms generally show substantial
inherent variability. The frequent occurrence of inversions in the concentration-effect
curves suggests that an overly specific or overly literal interpretation of each individual
data point might not be well founded. With regard to the exposure concentration
associated with the effect, uncertainties are introduced by the time variability of the
concentration of ammonia during the tests, and by longitudinal gradients in the streams
during some of the tests. Horizontal placement of points is subject to uncertainties
caused by the time variability of pH, and might be subject to a slightly low bias in some
cases. Finally, the elevated concentrations of ammonia yielded other changes (e.g.,
depressed concentration of dissolved oxygen) that confound the attribution of effects
solely to ammonia toxicity, although many of the data points appear to have little
potential to be affected by such other changes.
Some patterns can nevertheless be recognized in the data in Figure 15. Considering
the inherent variability, concentrations of ammonia below the 1998 CCC appear to yield
no significant effects relative to the control treatment. At concentrations above the
CCC applied as a 30-day average, many species experienced substantial stress,
although certain species might flourish under the conditions associated with such
concentrations of ammonia. Concentrations more than fourfold above the CCC applied
as a 30-day average appeared to yield conditions intolerable to many tested species.
At the tested temperatures, the 1999 CCC is somewhat higher than the 1998 CCC.
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Consequently, the 1999 CCC provides slightly less margin below the observed effect
levels shown on the log scale plot.
Tests with two species, the fathead minnow and the fingernail clam, occurred during a
time period when the pH was so variable that the CCC applied as a 4-day average was
substantially different than the CCC applied as a 30-day average. If applied simply as
a 30-day average, the CCC would have allowed substantial effects on the fingernail
clam. However, this species, which appeared to be the most sensitive tested species
in the study, would be protected by the additional limitation, which is expressed in the
1998 criterion statement, that the 4-day average concentration cannot be more than
two times the CCC (or 2.5 times in the 1999 criteria statement).
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Table 9. Data for Fishes and Clams in the Monticello Study8
Test
Fathead minnow
1st generation
Start 5/18/83
in lower reach
Fathead minnow
2nd generation
End 8/19/83
in lower reach
Bluegill
6/27/84-9/25/84
in lower reach
Channel catfish
1983
5/25/83-11/18/83
in lower reach
Channel catfish
1984A
5/7/84-6/12/84
in lower reach
Channel catfish
1984B
6/28/84-9/25/84
in lower reach
Duration
(Days)
63
63
90
177
36
89
Mean
Temp.
(C)
19.6
19.6
21.1
18.2
16.8
21.1
Mean Est. Max pH
PH
7.8
7.7
7.6
7.5
7.8
7.7
7.6
7.5
8.3
8.1
8.2
8.2
8.1
7.9
7.5
7.5
8.1
8.0
7.7
7.6
8.3
8.1
8.1
8.2
30-d
mean
7.8
7.7
7.6
7.5
7.8
7.7
7.6
7.5
8.3
8.1
8.2
8.2
8.5
8.4
8.0
8.0
8.1
8.0
7.7
7.6
8.3
8.1
8.1
8.2
4-d
mean
8.5
8.4
8.3
8.2
8.5
8.4
8.3
8.2
8.5
8.3
8.4
8.4
8.7
8.6
8.2
8.2
8.3
8.2
7.9
7.8
8.5
8.3
8.3
8.4
Est. Total Amm. N
'98 CCCb
(mg N/L)
1.14e
1.35e
1.59e
1.88e
1.14e
1.35e
1.59e
1.88e
0.80
1.10
0.94
0.94
0.57
0.67
1.27
1.27
1.10
1.27
1.87
2.08
0.80
1.10
1.10
0.94
Est. avg.
exp. cone
(mg N/L)
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
Rel. Conc.c
0.07
0.47
1.24
3.75
0.07
0.47
1.24
3.75
0.10
0.58
2.11
7.50
0.14
0.95
1.55
5.53
0.08
0.50
1.06
3.39
0.10
0.58
1.80
7.50
Biomass
Final
(g)
81
90
86
70
377
726
263
2437
1237
1489
1118
803
5138
4981
4385
3238
2108
2030
2202
1921
2923
2377
1204
1037
Rel.d
1.11
1.07
0.87
1.93
0.70
6.46
1.20
0.90
0.65
0.97
0.85
0.63
0.96
1.04
0.91
0.81
0.41
0.35
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White sucker 183 18.2
1983
5/19/83-11/18/83
in lower reach
White sucker 88 21.1
1984
6/29/84-9/25/84
in lower reach
Walleye yearling 46 24.1
6/29/84-8/14/84
in upper reach
Walleye young 43 16.7
8/20/84-10/2/84
in upper reach
Rainbow trout 237 5.9
1983-1984
10/19/83-6/12/84
in lower reach
Rainbow trout 69 10.6
1984A
9/6/84-11/14/84
in lower reach
Rainbow trout 28 5.9
1984B
10/16/84-11/13/84
in lower reach
8.4
7.9
7.5
7.5
8.3
8.1
8.1
8.2
8.2
8.1
8.0
8.2
8.4
8.3
8.4
8.4
8.3
8.1
7.8
7.7
8.3
8.2
8.1
8.4
8.1
7.9
8.1
8.4
8.7
8.4
8.0
8.0
8.3
8.1
8.1
8.2
8.2
8.1
8.0
8.2
8.4
8.3
8.4
8.4
8.6
8.4
8.1
8.0
8.3
8.2
8.1
8.4
8.1
7.9
8.2
8.4
8.8
8.6
8.2
8.2
8.5
8.3
8.3
8.4
8.4
8.3
8.2
8.4
8.5
8.4
8.5
8.5
8.6
8.4
8.1
8.0
8.5
8.4
8.3
8.6
8.3
8.1
8.3
8.6
0.41
0.67
1.27
1.27
0.80
1.10
1.10
0.94
0.94
1.10
1.27
0.94
0.67
0.80
0.67
0.67
0.48
0.67
1.10
1.27
0.80
0.94
1.10
0.67
1.10
1.46
1.10
0.67
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.08
0.64
1.98
7.04
0.20
0.95
1.55
5.53
0.10
0.58
1.80
7.50
0.09
0.58
1.55
7.50
0.12
0.80
2.93
10.44
0.17
0.95
1.80
5.53
0.10
0.68
1.80
10.44
0.08
0.44
1.80
10.44
2313
4287
3010
5854
4319
3866
3034
3366
2958
2731
2092
0
3056
2678
2178
0
5305
4514
5487
3630
1781
1971
948
0
403
420
252
201
1.85
1.30
2.53
0.90
0.70
0.78
0.92
0.71
0.00
0.88
0.71
0.00
0.85
1.03
0.68
1.11
0.53
0.00
1.04
0.63
0.50
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Fingernail clam A 7.9 8.7 0.81e 0.11 0.13 25f
6/6/83-8/1/83 7.9 8.5 1.14e 0.60 0.53 25f 1.01
7.9 8.6 0.96e 2.06 2.14 12f 0.48
7.9 8.5 1.14e 7.82 6.87 Of 0.00
Fingernail clam B 7.7 8.5 1.14e 0.11 0.09 llf
6/13/83-7/11/83 7.7 8.3 1.59e 0.60 0.38 14f 1.31
7.7 8.4 1.35e 2.06 1.53 2.4f 0.22
7.7 8.3 1.59e 7.82 4.91 Of 0.00
The data are from Hermanutz et al. (1987) and Zischke and Arthur (1987). All concentrations are total ammonia nitrogen and are
expressed as mg N/L.
The tabulated criterion is the lower of (1) the 1998 CCC calculated from the estimated maximum 30-day average pH or (2) two
times the 1998 CCC calculated from the estimated maximum 4-day average pH. Footnote e indicates where the latter condition
controlled the result.
Relative concentration = (treatment concentration/CCC calculated from the estimated maximum 30-day average pH).
Relative biomass = (treatment biomass/control biomass).
For the fathead minnow and the fingernail clam, two times the CCC calculated from the estimated maximum 4-day average pH
was less than the CCC calculated from the estimated maximum 30-day average pH.
For the fingernail clam, number of survivors times mean length is tabulated instead of biomass.
145
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Figure 15. Monticello data compared with the 1998 CCC statement
10 -i-
c
o
o
^
c
CD
"CD
CD
^ 1
(f>
CO
03
E
o
CO
CD
CD
01
0.1
fathead minnow /V bluegill
walleye | rainbow trout
channel catfish
fingernail clam
white sucker
Controls
I I III
I I III
0.01
0.1 1
Relative Cone. (treatment/CCC)
100
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Appendix 9. Water-Effect Ratios
Although the current guidance concerning Water-Effect Ratios (WERs) mainly
concerns their use with metals (U.S. EPA 1994), EPA provides for the determination
and use of WERs for ammonia. Because pH is the factor that has been shown to
substantially affect the toxicity of total ammonia in fresh water and the freshwater
criterion for ammonia is adjusted for pH, EPA expects that WERs for ammonia will
usually be close to 1. Indeed, most experimentally determined WERs for ammonia
have been close to 1:
a. Gersich and Hopkins (1986) and Mayes et al. (1986) reported that the acute and
chronic toxicity of ammonia in Tittabawassee River water was about the same as
reported by other investigators in laboratory dilution waters.
b. When Nimmo et al. (1989) compared a river water with a well water, the four WERs
ranged from 0.84 to 1.3; the four WERs obtained in comparisons of a wastewater
with the well water ranged from 0.5 to 1.5.
c. Diamond et al. (1993) obtained WERs of 1.1 and 2.0 with the fathead minnow and
Daphnia magna, respectively, using a well water and a pH-adjusted laboratory water.
d. In comparisons of a sewage effluent (pH=7.86 to 7.94) and a well water (pH=8.15 to
8.17), Monda et al. (1995) found WERs of 0.83 and 0.62 with a chironomid.
e. Using five species and waters from eight rivers, Willingham (1996) obtained
nineteen WERs that ranged from 0.57 to 1.47; one other WER was 3.
f. Acute and chronic tests with the fathead minnow and Ceriodaphnia dubia produced
four WERs that ranged from about 0.73 to 1.07 for Lake Mead (Willingham 1987).
g. Camp Dresser and McKee (1997) reported a WER of 2.5 with the fathead minnow,
but the test in site water lasted for seven days, whereas the tests in laboratory
dilution waters lasted for 30 and 350 days.
Although some of these WERs were not determined according to the guidance
presented in U.S. EPA (1984) and some might not have been adjusted for a pH
difference in the waters, they do illustrate that experimentally-determined WERs for
ammonia are likely to be close to 1.
It is possible that WERs for ammonia might be substantially different from 1 if there is
an interaction with other pollutants or if there is a substantial difference in ionic
composition, possibly in conjunction with a difference in pH or hardness (Ankley et al.
1995; Borgmann 1994; Borgmann and Borgmann 1997; Russo et al. 1988). WERs
might also be different from 1 if they are used to derive criteria for ammonia at pH<6.5
or pH>9.0. The pH of each of the waters used in the determination of the WERs given
above was between 7.3 and 8.7, except that pH was not reported by Willingham
(1996). Even though it appears that most WERs for ammonia will usually be close to
1.0, States and Tribes may nevertheless determine and use WERs to derive site-
specific criteria for ammonia. For this purpose EPA recommends the procedures of
U.S. EPA (1994).
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