United States Office of Water EPA 822-R-98-008
Environmental Protection 4304 August 1998
Agency
SERA 1998 Update
of Ambient
Water Quality
Criteria
for
Ammonia
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\ UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
* WASHINGTON, D.C. 20460
August 1998 OFFICE OF
WATER
Office of Science and Technology Policy Recommendations
The criteria recommendations provided here under Clean Water Act (CWA) Section 304(a)(l)
serve as guidance to States, Territories, and authorized Tribes in developing water quality
standards under CWA Section 303(c), used as a basis for controlling discharges or releases of
pollutants. The material provided in this document constitutes the Agency's current Section
304(a)(l) guidance, and will continue to serve as such until EPA publishes a revision.
Freshwater Ammonia Criteria Guidance
EPA prepared this guidance as a revision of its 1984/1985 and 1992 freshwater ammonia criteria.
This document revises (a) the pH and temperature relationship of the Criteria Maximum
Concentration, (CMC or acute criterion) based on re-evaluation of the data in the 1984 criteria
document, (b) the Criteria Continuous Concentration (CCC or chronic criterion), including its pH
and temperature relationship, based on new data in addition to what was available for the 1984
document, and (c) the averaging period applicable to the CCC. The document does not address,
and is not intended to modify (d) the averaging period applicable to the CMC, or (e) the
recommended frequencies for excursions of the CMC or CCC, which remain as set forth in the
1985 "Guidelines for Deriving...Criteria for the Protection of Aquatic Organisms...".
Cold-Season Risk Management Policy Recommendations
Because the costs of biological treatment of ammonia increase substantially as the water
temperature drops, establishing the cold-season ammonia concentrations necessary for protecting
aquatic life uses is of particular importance. Two factors affect the appropriateness of the update
document's CCC during cold seasons. First, with respect to chronic toxicity of ammonia to fish,
the most sensitive life stages are early life stages, which in many, but not all water bodies, do not
occur in during the cold season. Second, for the most sensitive invertebrates, the toxicity of
ammonia appears to decrease with decreasing temperature. For this reason, EPA has concluded
that under some circumstances the cold-season CCC could be relaxed somewhat, although setting
the appropriate criteria value involves uncertainties.
In light of the evidence available, EPA recommends the following risk management policies with
regard to cold-season ammonia criteria:
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While the cold-season ammonia criterion may in some cases be different than the criterion
applicable to other seasons, all periods of the year should be covered by some ammonia
criterion.
If a state can make a finding that identifies a time of year when no sensitive life stages of
any fish species are ordinarily present in numbers affecting the sustainability of
populations, the criterion applicable to that time of year may be set as much as 3-fold
higher than the criterion applicable to the remainder of the year. Baseline and subsequent
biological monitoring in accordance with currently available EPA guidance should be
conducted to assure that the integrity of the aquatic community being protected is
maintained when these higher cold-season concentrations are allowed.
If a state can demonstrate, based on rigorous baseline and subsequent instream biological
monitoring, that particular eco-regions can fully support beneficial fisheries uses, defined
by appropriate biological measures, under the cold-season concentration regimes
occurring at monitored sites in the eco-region, then the state may set the cold-season
criterion more than 3-fold higher than the applicable criterion to accord with the results of
such analysis. In judging the adequacy of the instream biological monitoring, EPA would
rely on its May 1996 guidance "Biological Criteria, Technical Guidance for Streams and
Small Rivers" (EPA 822-B-96-001) or later updates when they become available.
Endangered or Threatened Species Policy Recommendations
Because the criteria are generally designed to protect 95 percent of all fish and aquatic
invertebrate taxa, there remains a small possibility that the criteria will not protect all listed
endangered or threatened species. Consequently, EPA recommends the following:
In adopting ammonia criteria for specific water bodies, States and Tribes may need to
develop site-specific modifications of the criteria to protect listed endangered or
threatened species, where sufficient data exist indicating that endangered or threatened
species are more sensitive to a pollutant than the species upon which the criteria are based.
Such modifications may be accomplished using either of the following two procedures: (1)
If the CMC is greater than 0.5 times the Species Mean Acute Value for a listed threatened
or endangered species, or a surrogate for such species, obtained from flow-through,
measured-concentration tests, then the CMC should be reset equal to 0.5 times that
Species Mean Acute Value. (The empirical factor 0.5 converts from a 50 percent
lethality concentration to a minimal-lethality concentration.) If CCC is greater than the
Species Mean Chronic Value of a listed threatened or endangered species or surrogate,
then the CCC should be reset to that Species Mean Chronic Value. (2) The site-specific
criteria may be calculated using the recalculation procedure for site-specific modifications
described in Chapter 3 of the U.S. EPA Water Quality Standards Handbook, Second
Edition-Revised (1994).
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EPA encourages the submission of additional data relevant to the appropriateness of the guidance
contained in this document. Questions or comments may be directed to Charles Stephan, U.S.
EPA, 6201 Congdon Blvd., Duluth, MN 55804 (TEL: 218-529-5219; FAX: 218-529-5003) or
Charles Delos, U.S. EPA, Mail Code 4304, 401 M Street SW, Washington, DC 20460 (E-mail:
delos.charles@epamail.epa.gov).
Tudor T. Davies, Director
Office of Science and Technology
Office of Water
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1998 Update of
Ambient Water Quality Criteria for Ammonia
August 1998
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Washington, B.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)
concerning toxicity values that 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 update has been reviewed by the Mid-Continent Ecology
Division, Duluth, MN (Office of Research and Development) and the
Office of Science and Technology (Office of Water), U.S.
Environmental Protection Agency, and approved for publication.
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
ACKNOWLEDGMENT
This 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 or both to: Charles Stephan, U.S. EPA, 6201
Congdon Blvd., Duluth, MN 55804 (TEL: 218-529-5219)(FAX: 218-529-
5003) .
VI11
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CONTENTS
Page
Notices ii
Acknowledgment ii
Appendices iv
Tables v
Figures vi
Introduction 1
Overview of Ammonia Toxicology 2
Temperature-Dependence of Ammonia Toxicity 9
pH-Dependence of Ammonia Toxicity 21
Derivation of the New CMC 30
Review and Analysis of Chronic Data 38
Derivation of the New CCC 68
Cold-Weather Conditions 73
CCC Averaging Period 78
Water-Effect Ratios 81
A Field Study Relevant to the CCC 83
The National Criterion for Ammonia in Fresh Water 91
References 94
IX
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APPENDICES
Page
1 Review of Some Toxicity Tests 107
2 Methods for Regression Analysis of pH Data Ill
3 Conversion of Results of Toxicity Tests 114
4 Acute Values 116
5 Histopathological Effects 126
6 Results of Regression Analyses of Chronic Data 131
7 Acute-Chronic Ratios 144
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TABLES
Page
1 Ranked Genus Mean Acute Values 31
2 EC20s from Acceptable Chronic Tests 65
3 Data for Fishes and Clams in the Monticello Study .... 86
4 Results Obtained using Simulated Samples 113
5 Genus Mean Acute-Chronic Ratios 146
6 Ordered Genus Mean Acute-Chronic Ratios 148
XI
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FIGURES
Page
1 The effect of temperature on ammonia toxicity in terms of
un-ionized ammonia (DeGraeve et al. (1987) 12
2 The effect of temperature on acute ammonia toxicity in
terms of total ammonia 15
3 The effect of temperature on pH-adjusted acute ammonia
toxicity in terms of total ammonia 16
4 The effect of temperature on normalized acute ammonia
toxicity in terms of total ammonia 18
5 The effect of temperature on chronic ammonia lethality to
fathead minnows in terms of total ammonia (DeGraeve et al.
1987) 19
6 The effect of pH on acute ammonia toxicity in terms of
total ammonia 25
7 The effect of pH on normalized acute ammonia toxicity in
terms of total ammonia 27
8 The effect of pH on chronic ammonia toxicity in terms of
total ammonia 28
9 Ranked Genus Mean Acute Values (GMAVs) with Criterion
Maximum Concentrations (CMCs) 35
10 Acute LC50s used in criteria derivation in relationship to
Final Acute Values (FAVs) and Criterion Maximum
Concentrations (CMCs) 36
11 Ranked Genus Mean Chronic Values (GMCVs) with the
Criterion Continuous Concentration (CCC) 69
12 Chronic EC20s used in criteria derivation in relationship
to Criterion Continuous Concentrations (CCCs) 70
13 Monticello data compared with the new CCC statement ... 89
XII
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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 concerning aquatic life criteria for ammonia (Heber
and Ballentine 1992; U.S. EPA 1989,1996). Also, results of
additional toxicity tests on ammonia have been published since
1985, which could affect the freshwater criterion for ammonia.
The purpose of this 1998 Update is 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.
This 1998 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 1984/1985 ammonia criteria document 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 new CMC is
derived using these revised equations and the acute toxicity data
in the 1984/1985 criteria document. Then, new and old chronic
toxicity data are evaluated and used to derive a new CCC.
Finally, cold-weather conditions, the CCC averaging period,
water-effect ratios, and a field study relevant to the CCC are
discussed. This 1998 Update does not address (1) the CMC
averaging period, (2) the frequency of allowed exceedences, or
(3) field studies other than the one mentioned above. This 1998
Update addresses only the freshwater criterion for ammonia and
does 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.
Three unpublished manuscripts that were cited in the 1984/1985
criteria document have been published as Broderius et al. (1985),
Erickson (1985), and Thurston et al. (1986). West (1985) was
published as Arthur et al. (1987) .
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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)
[NHJ [H+]
K =
[NH+
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:
f
NH
(4)
NH, NH,+
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
Total Amm o n i a
6 7 8 9 10
PH
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
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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 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 permeabilities, 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.
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 of Ammonia
fl
o
-H
4-1
M O'1
4-1
fl
CD
O
fj
O 0.01
u
o
-H
X
Total Ammonia
PH
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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.
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 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 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.
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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 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
LC50s 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. 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
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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 KC1 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. (1981b). 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.
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
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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 its effects 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
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 of fish, 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
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temperature, which can also result in restrictive discharge
limitations.
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 criteria document, and more recent data
reported by Arthur et al. (1987), DeGraeve et al. (1987), Nimmo
et al. (1989), andKnoph (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
surprising because three other studies (Reinbold and Pescitelli
1982a; Thurston et al. 1983; DeGraeve et al. 1987) reported a
significant effect of temperature on the acute toxicity of un-
ionized ammonia to the fathead minnow. This discrepancy might be
due to other factors confounding temperature effects in the tests
by Arthur et al. (1987) because these tests were not conducted
10
-------
simultaneously; rather they were conducted during different
seasons. For five invertebrate species 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.
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 LC50s; 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 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/L at 6, 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, LC50s 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, LC50s expressed in terms of un-ionized ammonia increased
substantially with temperature.
Even with these additional data, the shape of the temperature
relationship is incompletely resolved and more research is
needed, especially regarding chronic toxicity and differences
among species. Nevertheless, the acute data for fishes
overwhelmingly indicate that ammonia toxicity, expressed in terms
of un-ionized ammonia, decreases with increasing temperature.
11
-------
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 logLC versus temperature.
O)
TO
'c
o
E
•a
a>
N
1
0.8
0.6
0.4
0.2
0.1
s
Fathead M in now
96-hr LC50
5 10 15 20 25 30
Temperature (C)
O)
I 0.8
I 0.6
1 °-4
C
o
I
C
Z>
0.2
0.1
Channel Catfish
96-hr LC50
5 10 15 20 25 30
Temperature (C)
) 0.6
I 0.4
o
E
•a
a>
.N
'c
o
I
C
Z>
0.2
0.1
0.08
0.06
Fathead M in now
30-day LC20
5 10 15 20 25 30
Temperature (C)
12
-------
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. An un-ionized ammonia criterion should continue to be lower at
lower temperatures, consistent with the observed temperature
dependence of ammonia toxicity to the most sensitive species,
i.e., fishes. The need for this is well established for the
CMC, based on the acute toxicity of ammonia to several species
of fish. 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. For the CCC, the only available
dataset concerns chronic mortality, and it supports a
relationship similar to that for acute toxicity.
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 raises 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. The expression and implementation of the
ammonia criterion would be considerably simplified if temperature
13
-------
was dropped as a modifying factor, which might be 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
this 1998 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 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
14
-------
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
100
80
60
40
30
20
d Minnow
et al. 1983)
80
60
40
30
20
15
10
Rainbow Trout
(Thurston and Russo 1983) 100
80
20
15
Channel Catfish
(Gary 1976)
30
20
15
8
6
4
Channel Catfish
(Colt & Tchobanoglous 1976) 40
30
- - - • • ^ji IP fT 9n
15
10
8
Rainbow Trout
- (Ministry of Tech 1968)
O
- o
600
400
300
200
150
100
80
Atlantic Salmon
(Knopf 1992)
10
20
30
10
20
30
10
20
30
10
20
30
10
20
30
10
20
30
O)
o
lO
O
40
30
20
15
10
8
Bluegil
(Roseboom & Richey 1977)
J I
80
60
40
30
20
15
10
Channel Catfish
60 h
(Roseboom & Richey 1977)
40
J I
30 -
(Reinbold & Pescitelli 1982a) 40
_ (Reinbold & Pescitelli 1982a)
0 10 20 30 0 10 20 30 0 10 20 30 0 10 20 30 0 10 20 30 0 10 20 30
60
40
30
20
15
10
Fathead Minnow
(Reinbold & Pescitelli 1982a) 100 —
10
20
30
Three-Spined Stickleback ,nn
r (Hazel et al. 1971)
Fathead Minnow
(DeGraeve et al. 1987)
20 30 0 10 20
TEMPERATURE (C)
400
300
30
100
80
60
Channel Catfish
(DeGraeve et al. 1987)
10
20
30
60
40
30
20
15
10
Fathead Minnow
(Nimmo et al. 1989)
10
20
30
15
-------
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 oo
80
60
40
30
20
Fathead Minnow
(Thurston et al. 1983)
O
1 0 20
30
Rainbow Trout
(Thurston and Russo 1983) 100
80
60
40
30
20
1 5
Channel Catfish
(Gary 1976)
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)
"<~> O
- ^ 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
1 0 20
30
1 0 20
30
1 0 20
30
O
10
O
40
30
20
1 5
1 0
8
6
Bluegill
(Roseboom & Richey 1977)
1 0 20
30
80
60
40
30
20
1 5
1 0
Channel Catfish
60
(Roseboom & Richey 1977)
40
30
20
1 5
1 0
Largemouth Bass
(Roseboom & Richey 1977)
40
30
20
1 5
1 0
8
6
Rainbow Trout
_ (Reinbold & Pescitelli 1982a)
40
20
1 5
1 0
Bluegill
(Reinbold & Pescitelli 1982a) 40
30 -
1 0 20
30
1 0 20
30
1 0 20
30
1 0 20
30
1 0 20
30
60
40
30
20
1 5
1 0
FatheadMinnow
(Reinbold & Pescitelli 1982a) 1 00
80
~ 60
40
30
20
1 5
Striped Bass
(Hazel et al. 1971)
300
Three-Spined Stickleback
(Hazel et al. 1971)
10 20
30
10 20
40
30
40
Fathead Minnow
(DeGraeve et al. 1987)
10 20 30 0 10 20
TEMPERATURE (C)
400
300
1 00
80
60
Channel Catfish
(DeGraeve et al. 1987)
30
10 20
30
60
40
30
20
1 5
1 0
Fathead Minnow
(Nimmo et al. 1989)
10 20
30
1 D
-------
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 Nimmo et al. (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, including temperature as a modifying factor for a total
ammonia criterion is not justified, and the criterion derived
below is based on the acute and chronic toxicity of total ammonia
without adjustment for test temperature. It is not argued that
total ammonia toxicity is absolutely constant with temperature or
that whatever temperature dependence exists is the same for all
life stages of all species, but rather it is argued that the
available data do not show temperature effects that are
sufficiently large or consistent enough to allow a worthwhile,
reliable temperature adjustment, either generically for all
species or for individual species. For invertebrates, it should
be noted that this update's assumption that temperature has 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. However, the
available data do not contradict either assumption. Fortunately,
most invertebrate species are resistant to the acute toxicity of
17
-------
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
0.8
0.6
0.4
Not pH Adjusted
2
1 .5
1
0.8
0.6
0.4
1 0
1 5
20 25 30
pH Adjusted
10 15 20
TEMPERATURE (C)
25
30
V ChannelCatfish
O Rainbow Trout
© Bluegill
F£] ChannelCatfish
A LargemouthBass
^ Rainbow Trout
<*> Bluegill
• Fathead Minnow
• Striped Bass
A Stickleback
^ Fathead Minnow
+ ChannelCatfish
O AtlanticSalmo n
© Johnny Darter
• Fathead Minnow
-------
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 logLC 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.
1 00
50
O)
o
CM
O
TO
'c
O
E
ro
•5
20
1 0
Data Not pH Adjusted
1 00 i-
50
20
1 0
Data Adjusted to pH 7.5
10 15 20 25
Temperature (C)
30
10 15 20 25
Temperature (C)
30
19
-------
ammonia, although some are sensitive to the chronic toxicity of
ammonia.
The amount of uncertainty in this approach 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 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. In contrast, the chronic data in Figure 5 suggest
that the total ammonia CCC should be about 20% lower at 5°C 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.
20
-------
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:
LIM LIM
LC50. = - 2 - + - i - (5)
t - -
where the subscript t denotes total ammonia, LIMH and LIML are
asymptotic (limiting) LC50s 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 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:
LC50. = LIMT
t I L/l
(6)
* '
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=
(LC50t/8) as the intercept parameter instead of LIML. Switching
from LIML to LC50t 8 requires use of a term that is the ratio
between LC50
t/
t 8
and LIM
L
LC50t =
LC5°t,8
R
+
1
R
1 +
1+10
PH-pHT
(7)
All three of the above model equations are equivalent, differing
only in the way in which the parameters are formulated.
22
-------
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:
LC50 =
LC5(D
R
1 +10
1 + 10
R
1+10
(8)
LC501 =
LC50 Q
t, 8
R
V nH -R
I
p TT
1 +
,PH-pHT
(9)
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
23
-------
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, 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 6 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, the
24
-------
Figure 6. The effect of pH on acute ammonia toxicity in terms of total ammonia. Symbols
denote LCSOs, 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)
400
200
° 100
O
< 40
Z 20
O
2 10
2
< 4
_i
< 2
Green Sunfish
(McCormick et al. 1984)
" Macrobracmum rosenber
(Armstrong et al. 1978)
Channel Catfish
(Tomasso et al. 1980)
Rainbow Trout
(Lloyd & Herbert 1960)
Chironomus tentans
(Schubauer-Berigan et al. 1995)
Hyalella azteca
(Ankley et al. 1995)
PH
25
-------
same as for temperature, 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 LC50tj8 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 6. The data points and the
common regression line from the pooled analysis are also plotted
together in Figure 7 by dividing each point by the LC50tj8 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 6). 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 LC50s. 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 LC50s, whereas others, such as Daphnia sp. and H. azteca
have lower than expected LC50s. 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.
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 8). 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 datapoints for the
invertebrate from overwhelming the data for the fish, datapoints
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
26
-------
Figure 7. 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 6.
o
lO
O
J5
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
CH Rainbow Trout
.A Coho Salmon
\7 Daphnia sp.
Sma llmouthBass
O G reen Sunfish
EH Ra inbow Trout
A Prawn Larvae
V ChannelCatfish
O White Perch
• Gu ppy
• ChannelCatfish
.A. Lumbriculus
^ C h ironom u s
• Hyalella
7 8
PH
27
-------
Figure 8. 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.
40
4 20
O)
cT 1 o
LJJ
I 4
^
< 2
2 1
S mallmouth Bass
(Broderius et al. 1 985)
Ceriodaphnia dubia
(Johnson 1 995)
1 00 i-
O)
40
10 20
c\j
O
LJJ
< 10
-z.
o
< 4
_i
O 2
1
PH
PH
-------
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 7, 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 = IAV I I °-°489 + ^ ) (10)
t t8 - -
0.0676 2.91
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 (Russo et al. 1988).
29
-------
DERIVATION OF THE NEW 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. However, some newer studies of acute toxicity
known to this effort were examined to determine whether new data
might materially affect the CMC. These studies include Ankley et
al. 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 (LC50s and EC50s) 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 (Table 1) using the procedure described in the
1985 Guidelines (U.S. EPA 1985b). (The same genera are in Table
1 in this 1998 Update as are in Table 3 in the 1984/1985 criteria
document and the SMAVs and GMAVs in both tables are based on the
test results in Table 1 in the criteria document. The GMAVs in
the two tables are different because (a) pH and temperature are
addressed differently in the two sets of calculations, (b) the
30
-------
Table 1. Ranked Genus Mean Acute Values
Genus Mean
Acute Value
Rank (mq N/La)
34 388.8
33 246.0
Species
Caddisfly,
Philarctus quaeris
Crayfish,
Orconectes immunis
Species Mean
Acute Value
(mq N/La)
388.8
1466.
Crayfish,
Orconectes nais
41.27
32
210.6
Isopod,
Asellus racovitzai
210.6
31
30
189.2
115.5
Mayfly,
Ephemerella grandis
Mayfly,
Callibaetis skokianus
189.2
175.6
29
113.2
Mayfly,
Callibaetis sp.
Beetle,
Stenelmis sexlineata
75.93
113.2
28
27
108.3
97.82
Amphipod, 108.3
Crangonyx pseudogracilis
Tubificid worm, 97.82
Tubifex tubifex
26
93.52
Snail,
Helisoma trivolvis
93.52
25
24
23
77.10
73.69
51.73
Stonefly,
Arcynopteryx parallela
Snail,
Physa gyrina
Mottled sculpin,
Cottus bairdi
77.10
73.69
51.73
22
51.06
Mosquitofish,
Gambusia affinis
51.06
21
20
43.55
38.11
Fathead minnow,
Pimephales promelas
White sucker,
Catostomus commersoni
43.55
45.82
31
-------
Genus Mean Species Mean
Acute Value Acute Value
Rank (mq N/La) Species (mq N/La)
Mountain sucker, 31.70
Catostomus platyrhynchus
19 36.82 Cladoceran, 35.76
Daphnia magna
Cladoceran, 37.91
Daphnia pulicaria
18 36.39 Brook trout, 36.39
Salvelinus fontinalis
17 35.65 Clam, 35.65
Musculium transversum
16 34.44 Channel catfish, 34.44
Ictalurus punctatus
15 33.99 Cladoceran, 33.99
Simocephalus vetulus
14 33.14 Guppy, 33.14
Poecilia reticulata
13 32.82 Flatworm, 32.82
Dendrocoelum lacteum
12 30.89 White perch, 30.89
Morone americana
11 26.97 Stoneroller, 26.97
Campostoma anomalum
10 26.50 Smallmouth bass, 35.07
Micropterus dolomieu
Largemouth bass, 20.03
Micropterus salmoides
9 26.11 Walleye, 26.11
Stizostedion vitreum
8 25.78 Cladoceran, 25.78
Ceriodaphnia acanthina
7 25.60 Red shiner, 45.65
Notropis lutrensis
Spotfin shiner, 19.51
Notropis spilopterus
Steelcolor shiner, 18.83
Notropis whipplei
32
-------
Genus Mean Species Mean
Acute Value Acute Value
Rank (mq N/La) Species (mq 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.
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).
33
-------
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 no temperature adjustment is used in either the
criteria document or this 1998 Update for invertebrates even
though Table 3 in the 1984/1985 criteria document is based on un-
ionized ammonia whereas Table 1 in this 1998 Update is based on
total ammonia nitrogen.) 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 11.23 mg N/L, and so the FAV is
lowered to this value, as per the 1985 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 AVt 8 in equation 10 results
in the following equation for expressing the CMC as a function of
pH:
™~ 0.275 39.0
CMC =
+ 107.204-pH 1 + 10pH-7.204
If the four genera (Oncorhynchus, Prosopium, Salmo, and
Salvelinus) in the family Salmonidae are excluded from the
dataset in Table 1, 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 10 gives the CMC as a function of pH:
™^ 0.411 58.4
CMC =
+ 107.204-pH 1 + 10pH-7.204
Figure 9 shows the ranked GMAVs, the CMC with salmonids present,
and the CMC with salmonids absent, all at pH=8 . The GMAVs
represent LC50s, whereas the CMCs represent concentrations that
are lethal to substantially less than 50 percent of the
individuals in either the fifth percentile genus or a sensitive
important species.
FAVs and CMCs are plotted in Figure 10, along with all of the
individual total ammonia acute values, unadjusted for pH, 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
34
-------
Figure 9. Ranked Genus Mean Acute Values (GMAVs) with Criterion
Maximum Concentrations (CMCs).
1000 7
CO
II
Q.
-i—•
03
D)
E
CD
100 -•
10
CMC salmonids absent
CMC salmonids present
Ranked Genera
35
-------
Figure 10. Acute LCSOs used in criteria derivation in
relationship to Final Acute Values (FAVs) and
Criterion Maximum Concentrations (CMCs).
1000
en
E
100 -
10 -
O
=AV - 5th Pet
«/o Salmonids
rAV - Adult
Rainbow TrouF
Slew CMC
«/o Salmonids
New CMC
w/ Salmonids
Old CMC
(10, 20 C)
Other Invertebrates
Salmonid Fishes
Nonsalmonid Fishes
7
\
8
\
9
PH
36
-------
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.
37
-------
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, all of the chronic data used in the
1984/1985 criteria document and newer chronic data known to the
authors or suggested by reviewers were reviewed and analyzed 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 this project
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 U.S. EPA (1986) .
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
38
-------
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
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:
39
-------
Mean (mq/L) Lower Limit (mq/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 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
40
-------
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 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
41
-------
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.
ion:
The regression model used was based on the logistic equat
T = ° (14)
1 + A-C B
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 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.
42
-------
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:
T =
(15)
ioo-p
]_Q
where logECp (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) :
T; = arcsin(t/T. /100) (16)
The regression equation was similarly transformed and the
parameter T0 was formulated to be the transformed effect.
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).
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 W^ 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-SE^ (17)
43
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where SEP is the SE of P as reported by the authors or
calculated as (P(1-P)/N)^, and SEW is the SE of W as reported
by the authors or calculated from their data.
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
44
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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. 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 pH=8 using the relationship between chronic toxicity
and pH derived above on the basis of Broderius et al. (1985) and
Johnson (1995). 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
2, 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 7.30 mg N/L when adjusted to pH=8.
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
45
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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 is 0.94 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 2 because results of field
tests are not used in the derivation of Final Chronic Values
(U.S. EPA 1985b).
The SMCV at pH=8 is <2.62 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 <2.62 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 is
19.8 mg N/L, which is the SMCV.
46
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Ceriodaphnia dubia
Willingham (1987) conducted a 7-day life-cycle test starting
with
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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 at pH=7.92 and 20.1°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 2 because neither survival nor
reproduction was studied.
Adjusted to pH=8, the respective EC20s are 15.1 and 19.4 mg
N/L. The SMCV for this species is 17.1 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
-------
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, 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
49
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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
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.
Carassius 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 Fl 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.
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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 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, and 30.2°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 2 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.
51
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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
the section below titled "A Field Study Relevant to the CCC"
and associated figures and table.) 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 2 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.
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 the
section below titled "A Field Study Relevant to the CCC" and
associated figures and table.) Two separate tests were
started with individuals whose average weight was 10 g and
52
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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 2 because
they are from a field study.
The value of ">4.79 mg N/L" is included in Table 2 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.
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
53
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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 2, 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 2.
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 2.
54
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Hermanutz et al. (1987) studied survival and growth of
juvenile channel catfish in experimental streams. (See the
section below titled "A Field Study Relevant to the CCC" and
associated figures and table.) 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 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 2 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 2, 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
55
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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 2.
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 2 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
56
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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 LC50s would be 22.0 and 18.7 mg N/L,
respectively, but LC20s would be lower than LC50s. 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 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.
57
-------
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 LC50s 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 LC50s that
were determined using fish whose average initial weights
were 0.7 to 10 g. The 35-day LC50s 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 LC50s 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 LC50s 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.
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
58
-------
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
(1968) 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 the
section below titled "A Field Study Relevant to the CCC" and
associated figures and table.) 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 2 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 2, 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.
59
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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 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 SMCV is <4.16 mg N/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 2 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
60
-------
reproduction. No data from this test are included in Table 2
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
was density-dependent. Although overflows resulted in loss of
fish 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.
61
-------
Too few fish died to allow calculation of an EC20. Neither of
these tests provided results that can be included in Table 2.
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 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 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
62
-------
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 2.
Hermanutz et al. (1987) studied survival and growth of the
juvenile bluegills in experimental streams. (See the section
below titled "A Field Study Relevant to the CCC" and
associated figures and table.) 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 2 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 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.
63
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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 the section
below titled "A Field Study Relevant to the CCC" and
associated figures and table.) 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 2
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 2. EC20s from Acceptable Chronic Tests'
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
Effectb
42-d Juv
Survival
42-d Juv
Survival
LC
Reproduction
7-d LC
Reproduction
7-d LC
Reproduction
21-d LC
Reproduction
21-d LC
Reproduction
10-wk LC
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
(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
(mg N/L)
7.30
0.94
19.8
14.6
11.6
15.1
19.4
<1.45
1.97
2.92
5.12
SMCVC
at pH=8
(mg N/L)
<2.62
19.8
13.0
17.1
<1.45
3.09
GMCVC
at pH=8
(mg N/L)
<2.62
16.0
17.1
<1.45
3.09
65
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Species
Catostomus
commersoni
Ictalurus
punctatus
Oncorhynchus
clarki
Oncorhynchus
my kiss
Oncorhynchus
nerka
Reference
Reinbold and
Pescitelli 1982a
Swigert and
Spacie 1983
Reinbold and
Pescitelli 1982a
Colt and
Tchobanoglous
1978
Thurston et al .
1978
Thurston et al .
1984b
Burkhalter and
Kaya 1977
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
(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
(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
(mg N/L)
>4.79
8.84
<4.16e
GMCVC
at pH=8
(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
(mg N/L)
5.84
5.61
1.85
9.61
8.62
8.18
1.54
EC20C
at pH=8
(mg N/L)
7.44
4.88
1.35
3.57
4.01
6.50
4.65
SMCVC
at pH=8
(mg N/L)
6.03
1.35
4.56
GMCVC
at pH=8
(mg N/L)
2.85
4.56
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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DERIVATION OF THE NEW CCC
Nine Genus Mean Chronic Values (GMCVs) are presented in Table 2.
The five lowest total ammonia nitrogen GMCVs at pH=8 are <1.45 mg
N/L for Hyalella, <2 . 62 mg N/L for Musculium, 2.85 mg N/L for
Lepomis, 3.09 mg N/L for Pimephales, and 4.56 mg N/L for
Micropterus. The more resistant genera with GMCVs greater than
4.7 mg N/L are Catostomus, Ictalurus, Ceriodaphnia, and Daphnia.
Although Table 2 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 FCV poses a risk to this genus.
Although Table 2 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, calculations based on the GMCVs
in Table 2 should adequately reflect the intent of the 1985
Guidelines. Use of the GMCVs for Hyalella, Musculium, Lepomis,
and Pimephales in the fifth percentile calculation procedure
described in the 1985 Guidelines results in a FCV of <1.27 mg N/L
at pH=8 . N=10 is used in this calculation because a GMCV for an
insect is assumed to be greater than 4.7 mg N/L. This FCV is a
"less than" value because the lowest two GMCVs are "less than"
values. Because no GMCV for a salmonid species is used in the
calculation of the FCV, it is not possible to calculate FCVs with
salmonids present and absent, as was done above for the FAV. The
CCC is set to 1.27 mg N/L at pH=8 . Figure 11 shows the ranked
GMCVs and the CCC, all at pH=8 .
Substitution of this CCC at pH=8 for CVt/8 in equation 11 results
in the following equation for expressing the new CCC as a
function of pH:
0.0858 3.70
CCC =
+ ]_Q7.688-pH -L +
This equation is plotted in Figure 12, along with the old CCC and
the EC20s from Table 2. The new CCC is near the old 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, the new
CCC corresponds to acute-chronic ratios of (14.4 mg N/L) /(I. 27 mg
N/L) = 11.3 using the calculated FAV when salmonids are present
(but not lowered to protect large rainbow trout) and (16.8 mg
N/L) /(I. 27 mg N/L) =13.2 using the FAV when salmonids are
absent. These are in the range of the ACRs that can be derived
68
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Figure 11. Ranked Genus Mean Chronic Values (GMCVs) with the
Criterion Continuous Concentration (CCC).
100
oo
ii
Q.
"CD
D)
E
O
^
CD
10 -•
CCC
Ranked Genera
69
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Figure 12. Chronic EC20s used in criteria derivation in
relationship to Criterion Continuous Concentrations
(CCCs).
50
20
10
en
E
0.5
0.2
0.1
O
New CCC
Old CCC, 10 C
Old CCC, 20 C
O
D
O
O
Musculium
Ceriodaphnia
Daphnia
Hyalella
Pimephales
Catostomus
Ictalurus
Oncorhynchus
Lepomis
Micropterus
I
7
8
PH
-------
from the EC20s in Table 2 (see Appendix 7). The ACR used to
calculate the old CCC was 13.5 (Heber and Ballentine 1992).
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. Any substantial increase in the CCC derived with the
procedures in this 1998 Update would require a higher GMCV for
Hyalella and a higher SMCV for the recreationally important
bluegill.
c. Because acutely resistant taxa are under-represented in the
chronic dataset in Table 2, it could be argued that n, the
number of genera used in the calculation of the FCV, 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 only raise the CCC to 1.37 mg N/L
at pH=8 (although then the CCC would be lowered to equal the
SMCV for the bluegill).
d. 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.
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, although other data
indicate that it might not be affected by such concentrations
and 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.
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 ten, (2) it
71
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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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COLD-WEATHER CONDITIONS
Dischargers that use biological treatment of ammonia are likely
to find it most difficult to meet water quality-based discharge
limits for ammonia when the temperature is the lowest. This has
raised questions about whether criteria based on toxicity tests
conducted mainly at warm temperatures appropriately define
concentrations that should be met under cold-weather conditions.
Considerable data indicate that toxicity of total ammonia does
not vary significantly with temperature, but this is based on a
few kinds of tests conducted with fishes. Furthermore, if
criteria are based on endpoints for invertebrates, there is a
question of whether the endpoints might in fact be less sensitive
at colder temperatures. Even if the toxicity of total ammonia is
independent of temperature for all endpoints, criteria should not
necessarily be independent of temperature unless the endpoints
upon which they are based are relevant during all portions 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-weather conditions (fish early life stages,
Hyalella reproduction) and in part on endpoints that might be
less sensitive under colder temperatures (fingernail clam
survival). Therefore, it is necessary to consider to what extent
and under what conditions the CCC can be higher during cold-
weather conditions.
An important consideration regarding raising the CCC during cold-
weather conditions is whether early life stages of fishes are
absent, which is not necessarily true for many waters. For
example, salmonids can spawn in cold temperatures in late fall or
early spring, so that early life stages can be present throughout
cold-weather conditions in such waters. Similarly, perch spawn
during cold-weather conditions in some waters, and early life
stages of some warmwater species are present during cold-weather
conditions in some southern waters. Furthermore, in some
situations, it might be necessary to limit the concentration of
ammonia in a discharge before spawning begins in order to ensure
that the concentration of ammonia is acceptably low at the site
soon enough in the reproductive cycle.
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Nevertheless, it is likely that there are bodies of water for
which some of the endpoints upon which the CCC is based are not
relevant during cold-weather conditions, and there is thus some
potential for the CCC to be raised. Unfortunately, a good
determination of how high the CCC can be in such situations is
not possible because few data are available concerning the
chronic sensitivities of the relevant life stages at the relevant
temperatures. The data that are needed are the results of
toxicity tests that are sufficiently long, are conducted at
appropriately low temperatures, and determine the effects of
ammonia on survival of life stages that are present during cold-
weather conditions.
In the absence of such data, however, there are ways in which
available data can be used to provide some indication of how
different the CCC can be during cold-weather conditions.
Fish
If it is assumed that the toxicity of total ammonia to fish is
independent of temperature for each endpoint, the CCC at cold
temperatures can be based on chronic tests conducted at warm
temperatures if the results are based on sensitive chronic
endpoints that are relevant during cold-weather conditions.
Therefore, when early life stages of fish are not present, the
best indication of what the CCC should be under cold-weather
conditions would be chronic survival tests, at any
temperature, with juvenile and adult fishes.
The only chronic survival tests conducted over a range of
temperatures are those of DeGraeve et al. (1987), which
studied 30-day survival of juvenile fathead minnows. When
expressed in terms of total ammonia and adjusted to pH=8, the
EC20s were 9.6, 12.6, 19.3, and 15.9 mg N/L at 6, 10, 25, and
30°C respectively. In the life-cycle fathead minnow test by
Thurston et al. (1986), parental generation mortality over
several months exposure at 24°C was not significant at 7 mg
N/L but exceeded 90% at 14 mg N/L, suggesting an EC20 close to
10 mg N/L for long-term survival. This result is somewhat
more sensitive than the warmwater tests by DeGraeve et al.
(1987), but is still less sensitive than the SMCV by about
three-fold.
However, in contrast to early life stages being more sensitive
than juvenile and adult fathead minnows, results obtained with
channel catfish by Colt and Tchobanoglous (1978) and Robinette
(1976) suggest that growth and survival of juveniles is as or
more sensitive than early life stages, based on the EC20s from
Swigert and Spacie (1983) and Reinbold and Pescitelli
74
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(1982a,c) in Table 2. Colt and Tchobanoglous (1978)
incompletely reported mortality data for juvenile channel
catfish, but the available information indicates that the EC20
at 28°C is between 8.7 and 9.9 mg N/L, when adjusted to pH=8.
For a 21-day exposure of juvenile bluegills at pH=7.3, Diamond
et al. (1993) reported 30% mortality at 64 mg N/L, which is 31
mg N/L when adjusted to pH=8. Although this might seem to
suggest considerable resistance relative to early life-stage
bluegills, this was a short test and the raw data indicate
that mortality was just starting during the last few days of
the test. The LC20 for more extended exposures would almost
certainly be no higher than half of this concentration, and
quite likely lower than that.
Although the absence of early life stages during cold-weather
conditions will generally not be an issue for salmonids, the
chronic sensitivities of juvenile and adult trout can be
useful in estimating what criteria should be in the absence of
early life stages. When exposures began after sensitive
embryo stages of rainbow trout, Solbe and Shurben (1989) did
not observe mortality significantly above control values until
26 mg N/L total ammonia (15 mg N/L adjusted to pH=8), at which
the control-corrected mortality was 30% after a 49-day
exposure. As discussed earlier, Broderius and Smith (1979)
reported 30% mortality of rainbow trout during a 30-day
exposure to 15.1 mg N/L (adjusted to pH=8). Based on three
tests by Thurston and Russo (1983) in which the concentration
of DO was always above 60 percent of saturation, the average
35-day LC50 for rainbow trout in the 0.6 to 10 g range is
26.5 mg N/L at pH=8. If the average slope of the chronic
regressions is used, this would correspond to an LC20 of about
15 mg N/L. For juvenile cutthroat trout, Thurston et al.
(1978) reported LC50s which averaged 19.7 mg N/L when adjusted
to pH=8, which would correspond to an LC20 of about 11 mg N/L.
The above data suggest that juveniles and adults of some fish
species have chronic LC20s in the range of 9 to 15 mg N/L (at
pH=8). This is in contrast to GMCVs in the range of 3 to 8 mg
N/L in Table 2. It should be noted, however, that most of the
juvenile and adult tests cited above were relatively short
compared to the duration of cold-weather conditions of
concern. Also, they do not address to what extent ammonia
effects that are not directly lethal will affect survival
under field conditions in which food availability and other
stresses are less favorable than in the laboratory (Lemly
1996), especially considering that ammonia is more persistent
and therefore more widespread during cold-weather conditions.
Furthermore, any cold-weather criterion derived from these
75
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data should lie below the lowest GMCV because of the small
number of genera with which tests have been conducted.
Therefore, a criterion on the order of 9 mg N/L at pH=8 would
not likely provide adequate protection. There is no clear
evidence for how much lower this number should be; setting a
cold-weather criterion must involve some site-specific risk
management considerations.
Invertebrates
Of the two chronically sensitive invertebrates, the fingernail
clam chronic value is already based on long-term survival of
juveniles so it is a relevant endpoint for cold-weather
conditions. For Hyalella, long-term survival is almost as
sensitive as reproduction, and the Hyalella GMCV based on
survival would be <«1.45 mg N/L. Therefore, the CCC would not
change. However, a few data are available concerning the
temperature-dependence of ammonia toxicity to invertebrates
and so there is a possibility that survival is less sensitive
under cold-weather conditions and that the CCC could
consequently be raised.
Based on toxicity tests by Arthur et al. (1987) during
different seasons, the 96-hr LC50 for the fingernail clam,
when expressed in terms of total ammonia nitrogen and adjusted
to pH=8, is a factor of 1.9 higher at 15°C than at 21°C, and a
factor of 2.7 higher at 5°C. For an amphipod (Crangonyx
pseudogracilis), Arthur et al. (1987) reported that LCSOs were
about 6-fold higher at 12 to 13°C and 8-fold higher at 4°C
than at 25°C. The effect of temperature on the rate of
biochemical processes might, however, affect the results of
acute (i.e., short-term) tests more than the results of
chronic (i.e., long-term) tests. Furthermore, these tests
might be confounded by effects other than temperature because
they were performed during different seasons. Nevertheless,
they still indicate that these invertebrates are more
resistant to ammonia at colder temperatures and/or during
colder seasons.
The above discussion is not intended to provide a definitive
value for relaxation of the CCC during cold-weather conditions,
but rather to indicate what types of data would be useful for
determining this and how much relaxation might conceivably occur.
The degree of relaxation is uncertain because the available data
do not directly address the endpoints of concern during long-term
exposures under cold-weather conditions. Deciding whether a
cold-weather CCC is justified and what the value should be is
highly site specific and the information provided here should be
considered to provide only suggestions as to how it might be
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derived. Careful consideration is needed regarding what data
here, and from other sources, are most relevant to the site in
question and what uncertainty factors should be applied. Until
more relevant data are available, application of available
information to development of a site-specific cold-weather CCC
requires a degree of risk management, after consideration of
biological and climatic conditions at the site, but
incorporating an explicit relationship concerning season or
temperature into the national criterion would require further
research.
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CCC 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. 1981a) . 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 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
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average can be, but based on these fathead minnow test results,
two-fold higher concentrations should pose little risk.
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 the
highest 4-day average concentration to two 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 2. 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
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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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WATER-EFFECT RATIOS
Although the current guidance concerning Water-Effect Ratios
(WERs) mainly concerns their use with metals (U.S. EPA 1994), the
U.S. EPA allows 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
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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, dischargers may
determine and use WERs to derive site-specific criteria for
ammonia whenever they want, as long as sufficient WERs are
determined in an acceptable manner (U.S. EPA 1994).
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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-
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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 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
CCC.
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
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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.
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 3 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 13, which show
the biological effect, relative to the control treatment, on the
vertical axis, and the exposure concentration, relative to the
new CCC of 1.27 mg N/L, on the horizontal axis.
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Table 3. Data for Fishes and Clams in the Monticello Study0
Test Duration Mean
(Days) temp.
(C)
Fathead minnow 63 19.6
1st generation
Start 5/18/83
in lower reach
Fathead minnow 63 19.6
2nd generation
End 8/19/83
in lower reach
Bluegill 90 21.1
6/27/84-9/25/84
in lower reach
Channel catfish 177 18.2
1983
5/25/83-11/18/83
in lower reach
Channel catfish 36 16.8
1984A
5/7/84-6/12/84
in lower reach
Channel catfish 89 21.1
1984B
6/28/84-9/25/84
in lower reach
Mean Est. Max pH
PH
Est. Total Ammonia N
Rel. Cone.
30-d
mean
7
7
7
7
7
7
7
7
8
8
8
8
8
8
8
8
8
8
7
7
8
8
8
.8
.7
.6
.5
.8
.7
.6
.5
.3
.1
.2
.2
.5
.4
.0
.0
.1
.0
.7
.6
.3
.1
.1
4-d
mean
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
8,
7,
7,
8,
8,
8,
.5
.4
.3
.2
.5
.4
.3
.2
.5
.3
.4
.4
.7
.6
.2
.2
.3
.2
.9
.8
.5
.3
.3
8.2
8.4
Criterion
(mg
1.
1.
1.
1.
1.
1.
1.
1.
0.
1.
0.
0.
0.
0.
1.
1.
1.
1.
1.
2.
0.
1.
1.
0.
N/L)
14e
35e
59e
88e
14e
35e
59e
88e
80
10
94
94
57
67
27
27
10
27
87
08
80
10
10
94
est ,
exp ,
(mg
0,
0,
1,
7,
0,
0,
1,
7,
0,
0,
1,
7,
0,
0,
1,
7,
0,
0,
1,
7,
0,
0,
1,
7,
. ave .
. cone
N/L)
.08
.64
.98
.04
.08
.64
.98
.04
.08
.64
.98
.04
.08
.64
.98
.04
.08
.64
.98
.04
.08
.64
.98
.04
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.'
1,
1,
0,
1,
0,
6,
1,
0,
0,
0,
0,
0,
0,
1,
0,
0,
0,
0,
.11
.07
.87
.93
.70
.46
.20
.90
.65
.97
.85
.63
.96
.04
.91
.81
.41
.35
86
-------
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
7
7
7
8
8
8
8
8
8
8
8
8
8
8
8
8
8
7
7
8
8
8
8
8
7
8
8
.4
.9
.5
.5
.3
.1
.1
.2
.2
.1
.0
.2
.4
.3
.4
.4
.3
.1
.8
.7
.3
.2
.1
.4
.1
.9
.1
.4
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
7
8
8
.7
.4
.0
.0
.3
.1
.1
.2
.2
.1
.0
.2
.4
.3
.4
.4
.6
.4
.1
.0
.3
.2
.1
.4
.1
.9
.2
.4
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
8
.8
.6
.2
.2
.5
.3
.3
.4
.4
.3
.2
.4
.5
.4
.5
.5
.6
.4
.1
.0
.5
.4
.3
.6
.3
.1
.3
.6
0
0
1
1
0
1
1
0
0
1
1
0
0
0
0
0
0
0
1
1
0
0
1
0
1
1
1
0
.41
.67
.27
.27
.80
.10
.10
.94
.94
.10
.27
.94
.67
.80
.67
.67
.48
.67
.10
.27
.80
.94
.10
.67
.10
.46
.10
.67
0.
0.
1.
7.
0.
0.
1.
7.
0.
0.
1.
7.
0.
0.
1.
7.
0.
0.
1.
7.
0.
0.
1.
7.
0.
0.
1.
7.
08
64
98
04
08
64
98
04
08
64
98
04
08
64
98
04
08
64
98
04
08
64
98
04
08
64
98
04
0
0
1
5
0
0
1
7
0
0
1
7
0
0
2
10
0
0
1
5
0
0
1
10
0
0
1
10
.20
.95
.55
.53
.10
.58
.80
.50
.09
.58
.55
.50
.12
.80
.93
.44
.17
.95
.80
.53
.10
.68
.80
.44
.08
.44
.80
.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.
1.
2.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
1.
0.
1.
0.
0.
1.
0.
0.
85
30
53
90
70
78
92
71
00
88
71
00
85
03
68
11
53
00
04
63
50
-------
Fingernail clam A
6/6/83-8/1/83
Fingernail clam B
6/13/83-7/11/83
7.9
7.9
7.9
7.9
7.7
7.7
7.7
7.7
i.7
i.5
1.6
1.5
1.5
1.3
1.4
1.3
0.81e
1.14s
0.96e
1.14e
1.14e
1.59e
1.35e
1.59e
0.11
0.60
2.06
7.82
0.11
0.60
2.06
7.82
0.13
0.53
2.14
6.87
0.09
0.38
1.53
4.91
25f
25f
12f
Of
14f
2.4f
Of
1
0
0
1
0
0
.01
.48
.00
.31
.22
.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 CCC calculated from the estimated maximum 30-day average
pH or (2) two times the 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.
-------
Figure 13. Monticello data compared with the new CCC statement
10 -i-
C
o
o
^
c
CD
E
"co
£ 1 --
c/j
ro
E
o
in
CD
CD
01
0.1
fathead minnow /\ bluegill
walleye | rainbow trout —I— fingernail clam
channel catfish white sucker
i i i Mini—i i i ii
0.01
0.1 1
Relative Cone. (treatment/CCC)
100
89
-------
Uncertainties exist in the vertical and horizontal locations of
points in Figure 13. 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 13. Considering the inherent variability, concentrations
of ammonia below the 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.
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 criterion statement, that the 4-day
average concentration cannot be more than two times the CCC.
90
-------
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 a very
sensitive species is important at a site, freshwater aquatic life
should be protected if both of the following conditions are
satisfied:
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 calculated using the following
equation:
0.275 39.0
CMC =
-^QpH-7.204
In situations where salmonids do not occur, the CMC may be
calculated using the following equation:
CMC=
2. 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 calculated using the following
equation:
0.0858 3.70
CCC =
107'688-PH
588
and the highest four-day average within the 30-day period does
not exceed twice the CCC.
The numeric values of the CMC with salmonids present and absent
and the CCC are:
pH CMC with CMC with CCC
salmonids salmonids
present absent
6.5 32.6 48.8 3.48
6.6 31.3 46.8 3.42
6.7 29.8 44.6 3.36
6.8 28.1 42.0 3.28
91
-------
6.9 26.2 39.1 3.19
7.0 24.1 36.1 3.08
7.1 22.0 32.8 2.96
7.2 19.7 29.5 2.81
7.3 17.5 26.2 2.65
7.4 15.4 23.0 2.47
7.5 13.3 19.9 2.28
7.6 11.4 17.0 2.07
7.7 9.65 14.4 1.87
7.8 8.11 12.1 1.66
7.9 6.77 10.1 1.46
8.0 5.62 8.40 1.27
8.1 4.64 6.95 1.09
8.2 3.83 5.72 0.935
8.3 3.15 4.71 0.795
8.4 2.59 3.88 0.673
8.5 2.14 3.20 0.568
8.6 1.77 2.65 0.480
8.7 1.47 2.20 0.406
8.8 1.23 1.84 0.345
8.9 1.04 1.56 0.295
9.0 0.885 1.32 0.254
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 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.
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, although other data
indicate that it might not be affected by such concentrations
and 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.
92
-------
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.
Because the chronic values for two of the four most chronically
sensitive genera are based on tests with early life stages of
fish, there is some uncertainty in applying the CCC during
conditions, such as during cold-weather conditions, when such
life stages are not present. Furthermore, although the data for
the two most sensitive genera (i.e., Hyalella and fingernail
clam) do not involve this life-stage issue, the acute toxicity
data for these taxa indicate that they probably become more
resistant to total ammonia as the temperature decreases.
Nevertheless, without exercising a degree of risk management that
is beyond the scope of this 1998 Update, the available data do
not allow a determination of how much higher the CCC could be
during a period during which the temperature is low and early
life stages of fishes are absent.
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 CCC averaging period presented above supersede
those given in previous guidance concerning the aquatic life
criterion for ammonia in fresh water. This 1998 Update does 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).
Because the ammonia criterion is a function of pH, calculation of
the appropriate weighted average pH is complicated. For some
purposes, calculation of an average pH can be avoided. For
example, if samples are obtained from a receiving water over a
period of time during which pH is not constant, the pH and the
concentration of total ammonia in each sample should be
determined. For each sample, the criterion should be determined
at the pH of the sample, and then the concentration of total
ammonia nitrogen in the sample should be divided by the criterion
to determine a quotient. If the geometric mean of the quotients
is less than 1 over an appropriate period of time, there is no
evidence that the criterion has been exceeded.
93
-------
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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.
Acute toxicitv at 20°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:
107
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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.
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.
108
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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.
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.
Bluegill 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 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.
109
-------
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.
110
-------
Appendix 2. Methods for Regression Analysis of pH Data
Analysis of the available data relating ammonia toxicity to pH
using Equations 8 and 9 requires recognition that, unlike usual
regression analysis with one response variable, two response
variables (i.e., LC50U and LC50J 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:
k k
z = 2^ 2^/*..v..
1=1 j =1
(22)
(x,,-)] [y -f (X,,,,-)]
.
1U
u = l
where k is the number of dependent variables, n is the number of
datapoints, ylu is the observed value for the dependent variable
i, and f(xlu,») is the model prediction of the value of the
dependent variable i. If correlation coefficients are zero,
Equation 22 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±js
should be minimized; this results in a formulation similar to
Equation 22, 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
111
-------
above, this assumption results in applying standard least squares
regression techniques to Equations 8 and 9.
For assumed parameter values LC508=1.0, pHT=7.5, R=0.01, and
•=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 at 0.25
intervals) datasets were run. The other factor was the true
correlation between the errors in logLC50u and logLC50i: 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 4.
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 log10LC50t 8 (corresponding to only
2.5% bias in the error for LC50t/8). 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 8 and 9, rather than adopting more
complicated methods.
112
-------
Table 4. Results Obtained using Simulated Samples
Parameter
True Value
pKT
7.5
log10R
-2.0
log10LC50t,8
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%
113
-------
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
this 1998 Update, 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 1998 Update:
Acute value (AV) = 1.40 mg NH3/L
pH = 8.20
Temperature = 18.0°C
Step 1.
Equation 3 in this 1998 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:
[NHJ
a_ = 10(pH-pK) = 0.0543369
Step 3.
The AV in terms of total ammonia is calculated as:
[NHJ
Total ammonia = [NHJ + [NHJ = [NHJ +
0.0543369
= 27.1652 mg total ammonia/L
Step 4.
The AV in terms of total ammonia nitrogen is calculated as
follows:
114
-------
Total ammonia nitrogen = (27.1652 mg total ammonia/L)(14/17;
= 22.3713 mg N/L.
Step 5.
The AV in terms of total ammonia nitrogen is converted from
pH=8.2 to pH=8 using equation 10 in this 1998 Update:
AVt/8 = (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 1 in this
update.
115
-------
Appendix 4. Acute Values'
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
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
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
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
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
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
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
116
-------
Crangonyx pseudogracilis
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
5.63
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
8.00
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
12.1
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
215.97
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
215.85
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
West 1985
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
117
-------
Oncorhynchus clarki
Oncorhynchus clarki
Oncorhynchus clarki
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
0.66
0.62
0.52
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
7.80
7.80
7.78
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
12.8
12.4
12.2
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
37.75
36.55
32.57
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
26.13
25.30
21.76
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
Thurston et al. 1978
Thurston et al. 1978
Thurston et al. 1978
Calamari et al. 1977, 1981
Calamari et al. 1977, 1981
Calamari et al. 1977, 1981
Calamari et al. 1977, 1981
BroderiusS Smith 1979
HoltS Malcolm 1979
DeGraeve et al. 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
118
-------
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
0.632
0.618
0.410
0.390
0.752
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
7.85
7.87
7.71
7.71
7.84
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
13.1
12.1
11.4
11.5
13.0
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
31.55
31.80
32.02
30.22
38.69
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
23.89
24.97
18.95
17.89
28.77
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
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
119
-------
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
0.552
0.580
0.484
0.297
0.327
0.289
0.262
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
111
7.86
7.88
7.69
7.74
7.76
7.66
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
13.6
10.2
10.0
10.7
10.4
10.0
9.80
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
31.81
35.31
28.60
25.62
25.76
22.44
25.95
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
20.89
27.23
22.87
14.66
16.05
14.47
14.12
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
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS Russo 1983
ThurstonS 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
120
-------
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
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
0.497
0.643
0.56
0.79
0.40
1.02
0.77
0.97
0.26
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
7.75
7.75
8.34
8.28
8.43
8.16
8.60
8.50
7.70
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
12.7
13.0
5.0
12.8
3.0
14.2
3.3
14.9
3.6
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
32.09
40.58
17.32
15.40
11.86
23.39
15.27
10.09
38.52
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
20.34
25.73
33.37
26.39
27.20
31.76
48.41
26.48
22.41
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
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
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
121
-------
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
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
Pimephales promelas
2.73
2.59
0.832
2.33
2.17
1.61
1.27
0.775
1.51
1.85
1.73
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
8.03
8.06
7.67
8.05
8.05
7.94
7.76
7.66
7.87
7.83
7.91
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
22.1
22.0
13.9
13.0
13.6
19.1
19.0
13.4
15.8
22.0
18.9
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
47.60
42.58
58.84
74.65
66.48
42.26
50.28
58.23
58.91
50.58
49.26
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
50.35
47.69
32.55
82.04
73.06
37.75
32.44
31.68
46.25
36.94
41.62
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
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
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
122
-------
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
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
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
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
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
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
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
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
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
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
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
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
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
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
Reinbold & Pescitelli 1982b
Reinbold & Pescitelli 1982b
Swigert & Spacie 1 983
West 1985
West 1985
West 1985
West 1985
West 1985
Wallenetal. 1957
Wallenetal. 1957
Wallenetal. 1957
Wallenetal. 1957
123
-------
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
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
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
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
7.22
7.45
7.45
6.00
8.00
7.84
8.28
6.61
7.20
7.72
8.69
111
111
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
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
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
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
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
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
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
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
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
Reinbold & Pescitelli 1982b
Reinbold & Pescitelli 1982b
Reinbold & Pescitelli 1982b
Reinbold & 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
124
-------
Micropterus dolomieu
Micropterus dolomieu
Micropterus salmoides
Micropterus salmoides
Etheostoma spectabile
Etheostoma spectabile
Stizostedion vitreum
Stizostedion vitreum
Stizostedion vitreum
Stizostedion vitreum
Cottus bairdi
1.20
1.78
1.0
1.7
0.90
1.07
0.85
0.52
1.10
0.51
1.39
7.74
8.71
7.96
8.04
8.40
8.10
8.08
7.90
7.70
8.30
8.02
22.3
22.3
22.0
28.0
21.0
22.0
18.2
3.7
11.1
19.0
12.4
39.30
7.56
20.48
19.59
7.65
16.12
17.43
48.37
89.93
6.12
49.83
24.49
29.33
18.99
21.12
16.55
19.49
20.29
40.12
52.33
10.91
51.73
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
ThurstonS 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 1 in this 1998 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.
125
-------
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
126
-------
effect concentrations were 0.11 mg NH3-N/L at pH=7.78 for the
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 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.
127
-------
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
(FjJ 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. 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 of fish 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
F! 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
128
-------
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.
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;
129
-------
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.
130
-------
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. 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.
131
-------
FINGERNAIL CLAM, 42-DAY JUV, ANDERSON ET AL. 1978
100
80
"co
'> 60
co
O 40
CD
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)
10 20
FINGERNAIL CLAM, 42-DAY JUV, SPARKS AND SANDUSKY 1981
CO
0)
100 r
80
60
O 40
CD
Q.
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
132
-------
CERIODAPHNIA ACANTHINA, LIFE CYCLE, MOUNT 1982
15
03
O)
6
"oo
CD
"co
CD
Q.
O)
O
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
I 25
CO
O)
° 9n
^ 20
to
CD
-------
CERIODAPHNIA DUBIA, LIFE CYCLE, NIMMO ET AL 1989
c
03
CT
O
"oo
CD
CO
CD
Q.
O)
O
15
12
EC20 = 15.2 mg N/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
CO
oo
CD
-------
DAPHNIA MAGNA, LIFE CYCLE, REINBOLD AND PESCITELLI 1982a
30
I 25
CO
O)
° 9n
^ 20
to
CD
15
O
i_
CD
Q.
O)
C
10
EC20 = 21.7mgN/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
13b
-------
FATHEAD MINNOW, LIFE CYCLE, THURSTON ET AL. 1986
60 h
50
CO
.c
"co
I
I
CD
Q_
40
30
20
10
EC20 = 1.97 mg N/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
5
to
CO 2.0
"O
o
CO
"CO 1-5
to
to
CO
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
136
-------
FATHEAD MINNOW, 28-DAY ELS, MAYES ET AL. 1986
03
0)
100 r
80
60
O 40
0)
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)
10 20
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
100 r
80
"03
'> 60
(J)
"c
O 40
CD
Q_
20
EC20 = 12.2 mg N/L (8.2-18.1)
T = 6.0°C
pH = 7.83
j i
j i i
0.1 0.2 0.5 1 2 5 10 20
Total Ammonia (mg N/L)
50 100
137
-------
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
100 r
80
"03
'> 60
05
O 40
0)
Q.
20
EC20 = 18.0mg N/L (5.4-60)
T= 10.0°C
pH = 7.73
J | L
J | L
0.1 0.2 0.5 1 2 5 10 20 50 100
Total Ammonia (mg N/L)
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
100 r
80
"03
'> 60
05
"c
O 40
CD
Q_
20
O1—I-
0.1
EC20 = 39 mg N/L (29-52)
T = 25.4°C
pH = 7.35
j i i
j i i
0.2 0.5 1 2 5 10 20
Total Ammonia (mg N/L)
j i
50 100
138
-------
FATHEAD MINNOW, 30-DAY JUVENILE, DEGRAEVE ET AL. 1987
03
100 r
80
60
c
O 40
Q.
20
EC20 = 35 mgN/L (17-72)
T = 30.2°C
pH = 7.19
—>-
0.1
j I
j I
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
03
T3
O
CO
-t->
03
03
E
o
m
EC20 = 11.5 mgN/L (9.7-13.6)
T = 26.9°C
pH = 7.76
j i
j i
j i
0.1 0.2 0.5 1 2 5 10
Total Ammonia (mg N/L)
20
139
-------
CHANNEL CATFISH, 30-DAY ELS, REINBOLD AND PESCITELLI 1982a
70 r
O) 60
03
T3
50
03
-t->
O)
40
30
EC20 = 12.2 mgN/L (4.3-28.9)
T = 25.8°C
pH = 7.80
I I
I I
I I
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
03
100 r
80
60
05
"c
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
140
-------
GREEN SUNFISH, 30-DAY ELS, MCCORMICK ET AL. 1984
4 r
O)
CO
CO _
03 2
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)
10 20
BLUEGILL, 30-DAY ELS, SMITH ET AL. 1984
2.5 r
2.0
5
CO
CO
03
1.5
0.5
0.0
0.2
EC20 = 1.85 mg N/L (1.42-2.42)
T = 22.5°C
pH = 7.76
0.5 1
10 20
Total Ammonia (mg N/L)
141
-------
SMALLMOUTH BASS, 32-DAY ELS, BRODERIUS ET AL. 1985
200 r
D) 150
co
Q
co
100
CD
o
50
EC20 = 9.61 mg N/L (6.59-14.02)
T = 22.3°C
pH = 6.60
0.5
2 5 10 20
Total Ammonia (mg N/L)
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 20
Total Ammonia (mg N/L)
50
142
-------
SMALLMOUTH BASS, 32-DAY ELS, BRODERIUS ET AL. 1985
150
D)
100
co
Q
CM
co
-l-»
co
CD
E 50
o
in
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)
100
co
Q
CN
CO
-i-»
co
co
E 50
o
in
EC20 = 1.54 mg N/L (1.25-1.89)
T = 22.3°C
pH = 8.68
0.05 0.1 0.2 0.5 1 2
Total Ammonia (mg N/L)
10
143
-------
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 2 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 5,
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 5).
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 LC50s from
three laboratories was only 27.2 to 51.5 mg N/L when adjusted to
pH=8.
Table 6 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
144
-------
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.
145
-------
Table 5. Genus Mean Acute-Chronic Ratios
Species Chronic Results Acute Results3 Adjusted to pH=8
M. transversum
C. acanthina
C. dubia
D. magna
H. azteca
P. promelas
C. commersoni
I. punctatus
L. cyanellus
Refb
1
2
3
4
5
6
7
8
9
10
11
7
10
7
12
Temp
23.5
21.8
24
26
25
19
20
25
24
25
24
18
26
25
22
.5
.0
.0
.8
.1
.0
.2
.1
.8
.6
.9
.4
.0
pH
8.15
7.80
7
8
7
8
7
7
8
7
8
8
7
8
7
.15
.57
.8
.45
.92
.94
.0
.82
.0
.32
.76
.16
.9
EC20C
5.82
1.23
44
5
15
7
21
<1
1
3
5
>2
11
5
5
.9
.80
.2
.37
.7
.58
.97
.73
.12
.9
.5
.84
.61
Temp
24
26
25
20
19
22
22
19
19
22
18
25
25
22
15
15
25
26
22
.0
.0
.0
.0
.7
.1
.0
.1
.0
.0
.9
.9
.6
.0
.0
.4
.7
.2
.4
pH
7
8
7
8
8
8
8
7
7
7
7
7
7
8
8
8
7
8
7
.06
.61
.8
.50
.34
.03
.06
.94
.76
.83
.91
.78
.8
.14
.16
.14
.8
.28
.7
LC50C
105
14
41
26
61
48
42
42
50
50
49
41
42
25
30
29
32
8
57
.8
.3
.4
.3
.6e
.6e
.3e
.4e
.6e
.3e
.0
.8
.2
.3f
.7f
.8
.6
m
EC20
7.30
0.94
19
14
11
15
19
<1
1
2
5
>4
8
7
4
.8
.1
.6
.1
.4
.45
.97
.92
.12
.79
.35
.44
.88
LC50
24
48
31
70
119
51
47
37
32
36
41
27
29
33
41
39
22
14
32
.4
.6
.5
.2
•
.5
.8
.7
.2
.7
.5
.2
.4
.1
.4
.0
.6
.8
.8
ACRd
1
3
2
4
6
20
9
6
<8
2
2
6
.2
.4
.7
.6
.1
.7
.7
.5
.4
.7
.0
.7
GMACR
1.9
5.3
10.9
<8.4
2.7
7.6
146
-------
L. macrochirus 13 22.5 7.76 1.85 21.7 7.6 44.2 1.35 21.4 15.9
M. dolomieu 14 22.3 6.60 9.61 22.3 6.53 371. 3.57 59.3 16.6 7.4
7.25 8.62 22.3 7.16 117. 4.01 30.4 7.6
7.83 8.18 22.3 7.74 39.5 6.50 24.4 3.8
8.68 1.54 22.3 8.71 7.43 4.65 29.3 6.3
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).
(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) Mayes et al. 1986;
(12) McCormick et al. 1984; (13) Smith et al. 1984; (14) Broderius et al. 1985.
Expressed as total ammonia nitrogen (mg N/L). Three digits are retained in intermediate
calculations to reduce roundoff error in subsequent calculations.
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.
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.
Reinbold and Pescitelli 1982b.
147
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Table 6. 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
14!
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