7.1 TYPES OF DILUTION WATER
7.1.1 The type of dilution water used in effluent toxicity
tests will depend largely on the objectives of the study.
7.1.1.1
If the objective of the test is to estimate the absolute
chronic toxicity of the effluent, a synthetic (standard) dilution
water is used. If the test organisms have been cultured in water
which is different from the test dilution water, a second set of
controls, using culture water, should be included in the
test.
7.1.1.2
If the objective of the test is to estimate the chronic
toxicity of the effluent in uncontaminated receiving water, the
test may be conducted using dilution water consisting of a single
grab sample of receiving water (if non-toxic), collected either
upstream and outside the influence of the outfall, or with other
uncontaminated natural water (ground or surface water) or standard
dilution water having approximately the same characteristics
(hardness, alkalinity, and conductivity) as the receiving water.
Seasonal variations in the quality of receiving waters may affect
effluent toxicity. Therefore, the pH, alkalinity, hardness, and
conductivity of receiving water samples should be determined before
each use. If the test organisms have been cultured in water which
is different from the test dilution water, a second set of
controls, using culture water, should be included in the
test.
7.1.1.3
If the objective of the test is to determine the additive
or mitigating effects of the discharge on already contaminated
receiving water, the test is performed using dilution water
consisting of receiving water collected immediately upstream or
outside the influence of the outfall. A second set of controls,
using culture water, should be included in the test.
7.1.2
An acceptable dilution water is one which is appropriate
for the objectives of the test; supports adequate performance of
the test organisms with respect to survival, growth, reproduction,
or other responses that may be measured in the test (i.e.,
consistently meets test acceptability criteria for control
responses); is consistent in quality; and does not contain
contaminants that could produce toxicity. Receiving waters,
synthetic waters, or synthetic waters adjusted to approximate
receiving water characteristics may be used for dilution provided
that the water meets the above listed qualifications for an
acceptable dilution water. USEPA (2000a) provides additional
guidance on selecting appropriate dilution waters.
7.1.3
When dual controls (one control using culture water and
one control using dilution water) are used (see Subsections 7.1.1.1
- 7.1.1.3 above), the dilution water control should be used to
determine test acceptability. It is also the dilution water control
that should be compared to effluent treatments in the calculation
and reporting of test results. The culture water control should be
used to evaluate the appropriateness of the dilution water source.
Significant differences between organism responses in culture water
and dilution water controls could indicate toxicity in the dilution
water and may suggest an alternative dilution water source. USEPA
(2000a) provides additional guidance on dual controls.
7.2
STANDARD, SYNTHETIC DILUTION WATER
7.2.1
Standard, synthetic dilution water is prepared with
deionized water and reagent grade chemicals or mineral water
(Tables 3 and 4). The source water for the deionizer can be ground
water or tap water.
7.2.2
DEIONIZED WATER USED TO PREPARE STANDARD, SYNTHETIC,
DILUTION WATER
7.2.2.1 Deionized water is obtained from a MILLIPORE® MILLI-Q®,
MILLIPORE® QPAK™2 or equivalent system. It is advisable to provide
a preconditioned (deionized) feed water by using a Culligan®,
Continental®, or equivalent system in front of the MILLIPORE®
System to extend the life of the MILLIPORE® cartridges (see Section
5, Facilities, Equipment, and Supplies).
7.2.2.2
The recommended order of the cartridges in a
four-cartridge deionizer (i.e., MILLI-Q® System or equivalent) is
(1) ion exchange, (2) ion exchange, (3) carbon, and (4) organic
cleanup (such as ORGANEX-Q®, or equivalent) followed by a final
bacteria filter. The QPAK™2 water system is a sealed system which
does not allow for the rearranging of the cartridges. However, the
final cartridge is an ORGANEX-Q® filter, followed by a final
bacteria filter. Commercial laboratories using this system have not
experienced any difficulty in using the water for culturing or
testing. Reference to the MILLI-Q® systems throughout the remainder
of the manual includes all MILLIPORE® or equivalent
systems.
7.2.3
STANDARD, SYNTHETIC FRESHWATER
7.2.3.1 To prepare 20 L of synthetic, moderately hard,
reconstituted water, use the reagent grade chemicals in Table 3 as
follows:
1.
Place 19 L of MILLI-Q®, or equivalent, water in a
properly cleaned plastic carboy.
2.
Add 1.20 g of MgSO4, 1.92 g NaHCO3, and 0.080g KCl to the
carboy.
3.
Aerate overnight.
4.
Add 1.20 g of CaSO4•2H20 to 1 L of MILLI-Q® or equivalent
deionized water in a separate flask. Stir on magnetic stirrer until
calcium sulfate is dissolved, add to the 19 L above, and mix
well.
5.
For Ceriodaphnia dubia culturing and testing, add
sufficient sodium selenate (Na2SeO4) to provide 2 mg selenium per
liter of final dilution water.
6.
Aerate the combined solution vigorously for an additional
24 h to dissolve the added chemicals and stabilize the
medium.
7.
The measured pH, hardness, etc., should be as listed in
Table 3.
TABLE 3. PREPARATION OF SYNTHETIC FRESHWATER USING REAGENT GRADE
CHEMICALS1
1
Taken in part from Marking and Dawson (1973).
2
Add reagent grade chemicals to deionized water.
3
Approximate equilibrium pH after 24 h of aeration.
4
Expressed as mg CaCO3/L.
7.2.3.2
If large volumes of synthetic reconstituted water will be
needed, it may be advisable to mix 1 L portions of concentrated
stock solutions of NaHCO3, MgSO4, and KCl for use in preparation of
the reconstituted waters.
7.2.3.3
To prepare 20 L of standard, synthetic, moderately hard,
reconstituted water, using mineral water such as PERRIER® Water, or
equivalent (Table 4), follow the instructions below.
1.
Place 16 L of MILLI-Q® or equivalent water in a properly
cleaned plastic carboy.
2.
Add 4 L of PERRIER® Water, or equivalent.
3.
Aerate vigorously for 24 h to stabilize the
medium.
4.
The measured pH, hardness and alkalinity of the aerated
water will be as indicated in Table 4.
5.
This synthetic water is referred to as diluted mineral
water (DMW) in the toxicity test methods.
TABLE 4. PREPARATION OF SYNTHETIC FRESHWATER USING MINERAL
WATER1
Approximate Final Water Quality
1
From Mount et al. (1987), and data provided by Philip Lewis,
EMSL-Cincinnati, OH.
2
Add mineral water to Milli-Q® water, or equivalent, to prepare
Diluted Mineral Water (DMW).
3
Approximate equilibrium pH after 24 h of aeration.
4
Expressed as mg CaCO3/L.
5
Dilutions of PERRIER® Water form a
precipitate when concentrations equivalent to "very hard water" are
aerated.
7.3 USE OF RECEIVING WATER AS DILUTION WATER
7.3.1
If the objectives of the test require the use of
uncontaminated receiving water as dilution water, and the receiving
water is uncontaminated, it may be possible to collect a sample of
the receiving water upstream of, or close to, but outside of the
zone influenced by the effluent. However, if the receiving water is
contaminated, it may be necessary to collect the sample in an area
"remote" from the discharge site, matching as closely as possible
the physical and chemical characteristics of the receiving water
near the outfall.
7.3.2
The sample should be collected immediately prior to the
test, but never more than 96 h before the test begins. Except where
it is used within 24 h, or in the case where large volumes are
required for flow through tests, the sample should be chilled to
0-6°C during or immediately following collection, and maintained at
that temperature prior to use in the test.
7.3.3
Receiving water containing debris or indigenous organisms
that may be confused with or attack the test organisms should be
filtered through a sieve having 60 mm mesh openings prior to
use.
7.3.4
Where toxicity-free dilution water is required in a test,
the water is considered acceptable if test organisms show the
required survival, growth, and reproduction in the controls during
the test.
7.3.5
The regulatory authority may require that the hardness of
the dilution water be comparable to the receiving water at the
discharge site. This requirement can be satisfied by collecting an
uncontaminated receiving water with a suitable hardness, or
adjusting the hardness of an otherwise suitable receiving water by
addition of reagents as indicated in Table 3.
7.4
USE OF TAP WATER AS DILUTION WATER
7.4.1
The use of tap water as dilution water is discouraged
unless it is dechlorinated and passed through a deionizer and
carbon filter. Tap water can be dechlorinated by deionization,
carbon filtration, or the use of sodium thiosulfate. Use of 3.6
mg/L (anhydrous) sodium thiosulfate will reduce 1.0 mg chlorine/L
(APHA, 1992). Following dechlorination, total residual chlorine
should not exceed 0.01 mg/L. Because of the possible toxicity of
thiosulfate to test organisms, a control lacking thiosulfate should
be included in toxicity tests utilizing thiosulfatedechlorinated
water.
7.4.2
To be adequate for general
laboratory use following dechlorination, the tap water is passed
through a deionizer and carbon filter to remove toxic metals and
organics, and to control hardness and alkalinity.
7.5
DILUTION WATER HOLDING
7.5.1 A given batch of dilution water should not be used for
more than 14 days following preparation because of the possible
build-up of bacterial, fungal, or algal slime growth and the
problems associated with it. The container should be kept covered
and the contents should be protected from light.
SECTION 8
EFFLUENT AND RECEIVING WATER SAMPLING, SAMPLE HANDLING, AND
SAMPLE PREPARATION FOR TOXICITY TESTS
8.1 EFFLUENT SAMPLING
8.1.1
The effluent sampling point should be the same as that
specified in the NPDES discharge permit (USEPA, 1988a). Conditions
for exception would be: (l) better access to a sampling point
between the final treatment and the discharge outfall; (2) if the
processed waste is chlorinated prior to discharge, it may also be
desirable to take samples prior to contact with the chlorine to
determine toxicity of the unchlorinated effluent; or (3) in the
event there is a desire to evaluate the toxicity of the influent to
municipal waste treatment plants or separate wastewater streams in
industrial facilities prior to their being combined with other
wastewater streams or non-contact cooling water, additional
sampling points may be chosen.
8.1.2
The decision on whether to collect grab or composite
samples is based on the objectives of the test and an understanding
of the short and long-term operations and schedules of the
discharger. If the effluent quality varies considerably with time,
which can occur where holding times are short, grab samples may
seem preferable because of the ease of collection and the potential
of observing peaks (spikes) in toxicity. However, the sampling
duration of a grab sample is so short that full characterization of
an effluent over a 24-h period would require a prohibitively large
number of separate samples and tests. Collection of a 24-h
composite sample, however, may dilute toxicity spikes, and average
the quality of the effluent over the sampling period. Sampling
recommendations are provided below (also see USEPA,
2002a).
8.1.3
Aeration during collection and transfer of effluents
should be minimized to reduce the loss of volatile
chemicals.
8.1.4
Details of date, time, location, duration, and procedures
used for effluent sample and dilution water collection should be
recorded.
8.2
EFFLUENT SAMPLE TYPES
8.2.1 The advantages and disadvantages of effluent grab and
composite samples are listed below:
8.2.1.1 GRAB SAMPLES Advantages:
1.
Easy to collect; require a minimum of equipment and
on-site time.
2.
Provide a measure of instantaneous toxicity. Toxicity
spikes are not masked by dilution. Disadvantages:
1. Samples are collected over a very short period of time and on
a relatively infrequent basis. The chances of detecting a spike in
toxicity would depend on the frequency of sampling and the
probability of missing a spike is high.
8.2.1.2 COMPOSITE SAMPLES Advantages:
1.
A single effluent sample is collected over a 24-h
period.
2.
The sample is collected over a much longer period of time
than a single grab sample and contains all toxicity
spikes.
Disadvantages:
1.
Sampling equipment is more sophisticated and expensive,
and must be placed on-site for at least 24 h.
2.
Toxicity spikes may not be detected because they are
masked by dilution with less toxic wastes.
8.3 EFFLUENT SAMPLING RECOMMENDATIONS
8.3.1
When tests are conducted on-site, test solutions can be
renewed daily with freshly collected samples, except for the green
alga, Selenastrum capricornutum, test which is not
renewed.
8.3.2
When tests are conducted off-site, a minimum of three
samples are collected. If these samples are collected on Test Days
1, 3, and 5, the first sample would be used for test initiation,
and for test solution renewal on Day 2. The second sample would be
used for test solution renewal on Days 3 and 4. The third sample
would be used for test solution renewal on Days 5, 6, and
7.
8.3.3
Sufficient sample volume must be collected to perform the
required toxicity and chemical tests. A 4-L (1gal) CUBITAINER® will
provide sufficient sample volume for most tests.
8.3.4
THE FOLLOWING EFFLUENT SAMPLING METHODS ARE
RECOMMENDED:
8.3.4.1 Continuous Discharges
8.3.4.1.1
If the facility discharge is continuous, a single 24-h
composite sample is to be taken.
8.3.4.2
Intermittent discharges
8.3.4.2.1
If the facility discharge is intermittent, a composite
sample is to be collected for the duration of the discharge but not
more than 24 hours.
8.4
RECEIVING WATER SAMPLING
8.4.1
Logistical problems and difficulty in securing sampling
equipment generally preclude the collection of composite receiving
water samples for toxicity tests. Therefore, based on the
requirements of the test, a single grab sample or daily grab sample
of receiving water is collected for use in the test.
8.4.2
The sampling point is determined by the objectives of the
test. In rivers, samples should be collected from mid-stream and at
mid-depth, if accessible. In lakes the samples are collected at
mid-depth.
8.4.3
To determine the extent of the zone of toxicity in the
receiving water downstream from the outfall, receiving water
samples are collected at several distances downstream from the
discharge. The time required for the effluent-receiving-water
mixture to travel to sampling points downstream from the outfall,
and the rate and degree of mixing, may be difficult to ascertain.
Therefore, it may not be possible to correlate downstream toxicity
with effluent toxicity at the discharge point unless a dye study is
performed. The toxicity of receiving water samples from five
stations downstream from the discharge point can be evaluated using
the same number of test vessels and test organisms as used in one
effluent toxicity test with five effluent dilutions.
8.5
EFFLUENT AND RECEIVING WATER SAMPLE HANDLING,
PRESERVATION, AND SHIPPING
8.5.1
Unless the samples are used in an on-site toxicity test
the day of collection (or hand delivered to the testing laboratory
for use on the day of collection), they should be chilled and
maintained at 0-6°C until used to inhibit microbial degradation,
chemical transformations, and loss of highly volatile toxic
substances.
8.5.2
Composite samples should be chilled as they are
collected. Grab samples should be chilled immediately following
collection.
8.5.3
If the effluent has been chlorinated, total residual
chlorine must be measured immediately following sample
collection.
8.5.4
Sample holding time begins when the last grab sample in a
series is taken (i.e., when a series of four grab samples are taken
over a 24-h period), or when a 24-h composite sampling period is
completed. If the data from the samples are to be acceptable for
use in the NPDES Program, the lapsed time (holding time) from
sample collection to first use of each grab or composite sample
must not exceed 36 h. EPA believes that 36 h is adequate time to
deliver the samples to the laboratories performing the test in most
cases. In the isolated cases, where the permittee can document that
this delivery time cannot be met, the permitting authority can
allow an option for onsite testing or a variance for an extension
of shipped sample holding time. The request for a variance in
sample holding time, directed to the USEPA Regional Administrator
under 40 CFR 136.3(e) should include supportive data which show
that the toxicity of the effluent sample is not reduced (e.g.,
because of volatilization and/or sorption of toxics on the sample
container surfaces) by extending the holding time beyond more than
36 h. However, in no case should more than 72 h elapse between
collection and first use of the sample. In static-renewal tests,
each grab or composite sample may also be used to prepare test
solutions for renewal at 24 h, 48 h, and/or 72 h after first use,
if stored at 0-6°C, with minimum head space, as described in
Subsection 8.5. If shipping problems (e.g., unsuccessful Saturday
delivery) are encountered with renewal samples after a test has
been initiated, the permitting authority may allow the continued
use of the most recently used sample for test renewal. Guidance for
determining the persistence of the sample is provided in Subsection
8.7.
8.5.5
To minimize the loss of toxicity due to volatilization of
toxic constituents, all sample containers should be "completely"
filled, leaving no air space between the contents and the
lid.
8.5.6
SAMPLES USED IN ON-SITE TESTS
8.5.6.1
Samples collected for on-site tests should be used within
24 h.
8.5.7
SAMPLES SHIPPED TO OFF-SITE FACILITIES
8.5.7.1
Samples collected for off-site toxicity testing are to be
chilled to 0-6°C during or immediately after collection, and
shipped iced to the performing laboratory. Sufficient ice should be
placed with the sample in the shipping container to ensure that ice
will still be present when the sample arrives at the laboratory and
is unpacked. Insulating material should not be placed between the
ice and the sample in the shipping container unless required to
prevent breakage of glass sample containers.
8.5.7.2
Samples may be shipped in one or more 4-L (l-gal)
CUBITAINERS® or new plastic "milk" jugs. All sample containers
should be rinsed with source water before being filled with sample.
After use with receiving water or effluents, CUBITAINERS® and
plastic jugs are punctured to prevent reuse.
8.5.7.3
Several sample shipping options are available, including
Express Mail, air express, bus, and courier service. Express Mail
is delivered seven days a week. Saturday and Sunday shipping and
receiving schedules of private carriers vary with the
carrier.
8.6
SAMPLE RECEIVING
8.6.1
Upon arrival at the laboratory, samples are logged in and
the temperature is measured and recorded. If the samples are not
immediately prepared for testing, they are stored at 0-6°C until
used.
8.6.2
Every effort must be made to initiate the test with an
effluent sample on the day of arrival in the laboratory, and the
sample holding time should not exceed 36 h unless a variance has
been granted by the NPDES permitting authority.
8.7
PERSISTENCE OF EFFLUENT TOXICITY DURING SAMPLE SHIPMENT
AND HOLDING
8.7.1
The persistence of the toxicity of
an effluent prior to its use in a toxicity test is of interest in
assessing the validity of toxicity test data, and in determining
the possible effects of allowing an extension of the holding time.
Where a variance in holding time (> 36 h, but # 72 h) is
requested by a permittee, (see Subsection 8.5.4 above), information
on the effects of the extension in holding time on the toxicity of
samples must be obtained by comparing the results of
multi-concentration chronic toxicity tests performed on effluent
samples held 36 h with toxicity test results using the same samples
after they were held for the requested, longer period. The portion
of the sample set aside for the second test should be held under
the same conditions as during shipment and holding.
8.8
PREPARATION OF EFFLUENT AND RECEIVING WATER SAMPLES FOR
TOXICITY TESTS
8.8.1
When aliquots are removed from the sample container, the
head space above the remaining sample should be held to a minimum.
Air which enters a container upon removal of sample should be
expelled by compressing the container before reclosing, if possible
(i.e., where a CUBITAINER® is used), or by using an appropriate
discharge valve (spigot).
8.8.2
With the daphnid, Ceriodaphnia dubia, and fathead minnow,
Pimephales promelas, tests, effluents and receiving waters should
be filtered through a 60-µm plankton net to remove indigenous
organisms that may attack or be confused with the test organisms
(see the daphnid, Ceriodaphnia dubia, test method for details).
Receiving waters used in green alga, Selenastrum capricornutum,
toxicity tests must be filtered through a 0.45-µm pore diameter
filter before use. It may be necessary to first coarse-filter the
dilution and/or waste water through a nylon sieve having 2- to 4-mm
mesh openings to remove debris and/or break up large floating or
suspended solids. Because filtration may increase the dissolved
oxygen (DO) in the effluent, the DO should be checked both before
and after filtering. Low dissolved oxygen concentrations will
indicate a potential problem in performing the test. Caution:
filtration may remove some toxicity.
8.8.3
If the samples must be warmed to bring them to the
prescribed test temperature, supersaturation of the dissolved
oxygen and nitrogen may become a problem. To avoid this problem,
samples may be warmed slowly in open test containers. If DO is
still above 100% saturation after warming to test temperature,
samples should be aerated moderately (approximately 500 mL/min) for
a few minutes using an airstone. If DO is below 4.0 mg/L after
warming to test temperature, the solutions must be aerated
moderately (approximately 500 mL/min) for a few minutes, using an
airstone, until the DO is within the prescribed range ($4.0 mg/L).
Caution: avoid excessive aeration.
8.8.4
The DO concentration in the samples should be near
saturation prior to use. Aeration may be used to bring the DO and
other gases into equilibrium with air, minimize oxygen demand, and
stabilize the pH. However, aeration during collection, transfer,
and preparation of samples should be minimized to reduce the loss
of volatile chemicals.
8.8.4.1
Aeration during the test may alter the results and should
be used only as a last resort to maintain the required DO. Aeration
can reduce the apparent toxicity of the test solutions by stripping
them of highly volatile toxic substances, or increase their
toxicity by altering pH. However, the DO in the test solutions
should not be allowed to fall below 4.0 mg/L.
8.8.4.2
In static tests (renewal or non-renewal), low DOs may
commonly occur in the higher concentrations of wastewater. Aeration
is accomplished by bubbling air through a pipet at a rate of 100
bubbles/min. If aeration is necessary, all test solutions must be
aerated. It is advisable to monitor the DO closely during the first
few hours of the test. Samples with a potential DO problem
generally show a downward trend in DO within 4 to 8 h after the
test is started. Unless aeration is initiated during the first 8 h
of the test, the DO may be exhausted during an unattended period,
thereby invalidating the test.
8.8.5
At a minimum, pH, conductivity, and total residual
chlorine are measured in the undiluted effluent or receiving water,
and pH and conductivity are measured in the dilution
water.
8.8.5.1
It is recommended that total alkalinity and total
hardness also be measured in the undiluted effluent test water,
receiving water, and the dilution water.
8.8.6
Total ammonia is measured in effluent and receiving water
samples where toxicity may be contributed by unionized ammonia
(i.e., where total ammonia $ 5 mg/L). The concentration (mg/L) of
un-ionized (free) ammonia in a sample is a function of temperature
and pH, and is calculated using the percentage value obtained from
Table 5, under the appropriate pH and temperature, and multiplying
it by the concentration (mg/L) of total ammonia in the
sample.
8.8.7
Effluents and receiving waters can be dechlorinated using
6.7 mg/L anhydrous sodium thiosulfate to reduce 1 mg/L chlorine
(APHA, 1992). Note that the amount of thiosulfate required to
dechlorinate effluents is greater than the amount needed to
dechlorinate tap water (see Section 7, Dilution Water, Subsection
7.4.1). Since thiosulfate may contribute to sample toxicity, a
thiosulfate control should be used in the test in addition to the
normal dilution water control.
8.8.8
Mortality or impairment of growth or reproduction due to
pH alone may occur if the pH of the sample falls outside the range
of 6.0 - 9.0. Thus, the presence of other forms of toxicity (metals
and organics) in the sample may be masked by the toxic effects of
low or high pH. The question about the presence of other toxicants
can be answered only by performing two parallel tests, one with an
adjusted pH, and one without an adjusted pH. Freshwater samples are
adjusted to pH 7.0 by adding 1N NaOH or 1N HCl dropwise, as
required, being careful to avoid overadjustment.
1
Table provided by Teresa Norberg-King, ERL, Duluth, Minnesota.
Also see Emerson et al. (1975), Thurston et al. (1974), and USEPA
(1985a).
8.9 PRELIMINARY TOXICITY RANGE-FINDING TESTS
8.9.1
USEPA Regional and State personnel generally have
observed that it is not necessary to conduct a toxicity
range-finding test prior to initiating a static, chronic,
definitive toxicity test. However, when preparing to perform a
static test with a sample of completely unknown quality, or before
initiating a flow-through test, it is advisable to conduct a
preliminary toxicity range-finding test.
8.9.2
A toxicity range-finding test ordinarily consists of a
down-scaled, abbreviated static acute test in which groups of five
organisms are exposed to several widely-spaced sample dilutions in
a logarithmic series, such as
100%, 10.0%, 1.00%, and 0.100%, and a control, for 8-24 h.
Caution: if the sample must also be used for the fullscale
definitive test, the 36-h limit on holding time (see Subsection
8.5.4) must not be exceeded before the definitive test is
initiated.
8.9.3 It should be noted that the toxicity
(LC50) of a sample observed in a range-finding test may be
significantly different from the toxicity observed in the follow-up
chronic definitive test because: (1) the definitive test is longer;
and (2) the test may be performed with a sample collected at a
different time, and possibly differing significantly in the level
of toxicity.
8.10 MULTI-CONCENTRATION (DEFINITIVE) EFFLUENT TOXICITY
TESTS
8.10.1
The tests recommended for use in determining discharge
permit compliance in the NPDES program are multi-concentration, or
definitive, tests which provide (1) a point estimate of effluent
toxicity in terms of an IC25, IC50, or LC50, or (2) a
no-observed-effect-concentration (NOEC) defined in terms of
mortality, growth, reproduction, and/or teratogenicity and obtained
by hypothesis testing. The tests may be static renewal or static
non-renewal.
8.10.2
The tests consist of a control and a minimum of five
effluent concentrations. USEPA recommends the use of a $0.5
dilution factor for selecting effluent test concentrations.
Effluent test concentrations of 6.25%, 12.5%, 25%, 50%, and 100%
are commonly used, however, test concentrations should be selected
independently for each test based on the objective of the study,
the expected range of toxicity, the receiving water concentration,
and any available historical testing information on the effluent.
USEPA (2000a) provides additional guidance on choosing appropriate
test concentrations.
8.10.3
When these tests are used in determining compliance with
permit limits, effluent test concentrations should be selected to
bracket the receiving water concentration. This may be achieved by
selecting effluent test concentrations in the following manner: (1)
100% effluent, (2) [RWC + 100]/2, (3) RWC, (4) RWC/2, and (5)
RWC/4. For example, where the RWC = 50%, appropriate effluent
concentrations may be 100%, 75%, 50%, 25%, and 12.5%.
8.10.4
If acute/chronic ratios are to be
determined by simultaneous acute and short-term chronic tests with
a single species, using the same sample, both types of tests must
use the same test conditions, i.e., pH, temperature, water
hardness, salinity, etc.
8.11
RECEIVING WATER TESTS
8.11.1
Receiving water toxicity tests generally consist of 100%
receiving water and a control. The total hardness of the control
should be comparable to the receiving water.
8.11.2
The data from the two treatments are analyzed by
hypothesis testing to determine if test organism survival in the
receiving water differs significantly from the control. Four
replicates and 10 organisms per replicate are required for each
treatment (see Summary of Test Conditions and Test Acceptability
Criteria in the specific test method).
8.11.3
In cases where the objective of the test is to estimate
the degree of toxicity of the receiving water, a multiconcentration
test is performed by preparing dilutions of the receiving water,
using a $ 0.5 dilution series, with a suitable control
water.
SECTION 9
CHRONIC TOXICITY TEST ENDPOINTS AND DATA
ANALYSIS
9.1 ENDPOINTS
9.1.1 The objective of chronic aquatic toxicity tests with
effluents and pure compounds is to estimate the highest "safe" or
"no-effect concentration" of these substances. For practical
reasons, the responses observed in these tests are usually limited
to hatchability, gross morphological abnormalities, survival,
growth, and reproduction, and the results of the tests are usually
expressed in terms of the highest toxicant concentration that has
no statistically significant observed effect on these responses,
when compared to the controls. The terms currently used to define
the endpoints employed in the rapid, chronic and sub-chronic
toxicity tests have been derived from the terms previously used for
full life-cycle tests. As shorter chronic tests were developed, it
became common practice to apply the same terminology to the
endpoints. The terms used in this manual are as follows:
9.1.1.1
Safe Concentration - The highest concentration of
toxicant that will permit normal propagation of fish and other
aquatic life in receiving waters. The concept of a "safe
concentration" is a biological concept, whereas the
"no-observed-effect concentration" (below) is a statistically
defined concentration.
9.1.1.2
No-Observed-Effect-Concentration (NOEC) - The highest
concentration of toxicant to which organisms are exposed in a full
life-cycle or partial life-cycle (short-term) test, that causes no
observable adverse effects on the test organisms (i.e., the highest
concentration of toxicant in which the values for the observed
responses are not statistically significantly different from the
controls). This value is used, along with other factors, to
determine toxicity limits in permits.
9.1.1.3
Lowest-Observed-Effect-Concentration (LOEC) - The lowest
concentration of toxicant to which organisms are exposed in a
life-cycle or partial life-cycle (short-term) test, which causes
adverse effects on the test organisms (i.e., where the values for
the observed responses are statistically significantly different
from the controls).
9.1.1.4
Effective Concentration (EC) - A point estimate of the
toxicant concentration that would cause an observable adverse
affect on a quantal, "all or nothing," response (such as death,
immobilization, or serious incapacitation) in a given percent of
the organisms, calculated by point estimation techniques. If the
observable effect is death or immobility, the term, Lethal
Concentration (LC), should be used (see Subsection 9.1.1.5). A
certain EC or LC value might be judged from a biological standpoint
to represent a threshold concentration, or lowest concentration
that would cause an adverse effect on the observed
response.
9.1.1.5
Lethal Concentration (LC) - The toxicant concentration
that would cause death in a given percent of the test population.
Identical to EC when the observed adverse effect is death. For
example, the LC50 is the concentration of toxicant that would cause
death in 50% of the test population.
9.1.1.6
Inhibition Concentration (IC) - The toxicant
concentration that would cause a given percent reduction in a
non-quantal biological measurement for the test population. For
example, the IC25 is the concentration of toxicant that would cause
a 25% reduction in mean young per female or in growth for the test
population, and the IC50 is the concentration of toxicant that
would cause a 50% reduction.
9.2
RELATIONSHIP BETWEEN ENDPOINTS DETERMINED BY HYPOTHESIS
TESTING AND POINT ESTIMATION TECHNIQUES
9.2.1 If the objective of chronic aquatic toxicity tests with
effluents and pure compounds is to estimate the highest "safe or
no-effect concentration" of these substances, it is imperative to
understand how the statistical endpoints of these tests are related
to the "safe" or "no-effect" concentration. NOECs and LOECs are
determined by hypothesis testing (Dunnett's Test, a t test with the
Bonferroni adjustment, Steel's Many-one Rank Test, or the Wilcoxon
Rank Sum Test with the Bonferroni adjustment), whereas LCs, ICs,
and ECs are determined by point estimation techniques (Probit
Analysis, Spearman-Karber Method, Trimmed Spearman-Karber Method,
Graphical Method or Linear Interpolation Method). There are
inherent differences between the use of a NOEC or LOEC derived from
hypothesis testing to estimate a "safe" concentration, and the use
of a LC, EC, IC, or other point estimates derived from curve
fitting, interpolation, etc.
9.2.2
Most point estimates, such as the LC, IC, or EC, are
derived from a mathematical model that assumes a continuous
dose-response relationship. By definition, any LC, IC, or EC value
is an estimate of some amount of adverse effect. Thus the
assessment of a "safe" concentration must be made from a biological
standpoint rather than with a statistical test. In this instance,
the biologist must determine some amount of adverse effect that is
deemed to be "safe", in the sense that from a practical biological
viewpoint it will not affect the normal propagation of fish and
other aquatic life in receiving waters.
9.2.3
The use of NOECs and LOECs, on the other hand, assumes
either (1) a continuous dose-response relationship, or (2) a
non-continuous (threshold) model of the dose-response
relationship.
9.2.3.1
In the case of a continuous dose-response relationship,
it is also assumed that adverse effects that are not "statistically
observable" are also not important from a biological standpoint,
since they are not pronounced enough to test as statistically
significant against some measure of the natural variability of the
responses.
9.2.3.2
In the case of non-continuous dose-response
relationships, it is assumed that there exists a true threshold, or
concentration below which there is no adverse effect on aquatic
life, and above which there is an adverse effect. The purpose of
the statistical analysis in this case is to estimate as closely as
possible where that threshold lies.
9.2.3.3
In either case, it is important to realize that the
amount of adverse effect that is statistically observable (LOEC) or
not observable (NOEC) is highly dependent on all aspects of the
experimental design, such as the number of concentrations of
toxicant, number of replicates per concentration, number of
organisms per replicate, and use of randomization. Other factors
that affect the sensitivity of the test include the choice of
statistical analysis, the choice of an alpha level, and the amount
of variability between responses at a given
concentration.
9.2.3.4
Where the assumption of a continuous dose-response
relationship is made, by definition some amount of adverse effect
might be present at the NOEC, but is not great enough to be
detected by hypothesis testing.
9.2.3.5
Where the assumption of a non-continuous dose-response
relationship is made, the NOEC would indeed be an estimate of a
"safe" or "no-effect" concentration if the amount of adverse effect
that appears at the threshold is great enough to test as
statistically significantly different from the controls in the face
of all aspects of the experimental design mentioned above. If,
however, the amount of adverse effect at the threshold were not
great enough to test as statistically different, some amount of
adverse effect might be present at the NOEC. In any case, the
estimate of the NOEC with hypothesis testing is always dependent on
the aspects of the experimental design mentioned above. For this
reason, the reporting and examination of some measure of the
sensitivity of the test (either the minimum significant difference
or the percent change from the control that this minimum difference
represents) is extremely important.
9.2.4
In summary, the assessment of a "safe" or "no-effect"
concentration cannot be made from the results of statistical
analysis alone, unless (1) the assumptions of a strict threshold
model are accepted, and (2) it is assumed that the amount of
adverse effect present at the threshold is statistically detectable
by hypothesis testing. In this case, estimates obtained from a
statistical analysis are indeed estimates of a "no-effect"
concentration. If the assumptions are not deemed tenable, then
estimates from a statistical analysis can only be used in
conjunction with an assessment from a biological standpoint of what
magnitude of adverse effect constitutes a "safe" concentration. In
this instance, a "safe" concentration is not necessarily a truly
"no-effect" concentration, but rather a concentration at which the
effects are judged to be of no biological significance.
9.2.5
A better understanding of the relationship between
endpoints derived by hypothesis testing (NOECs) and point
estimation techniques (LCs, ICs, and ECs) would be very helpful in
choosing methods of data analysis. Norberg-King (1991) reported
that the IC25s were comparable to the NOECs for 23 effluent and
reference toxicant data sets analyzed. The data sets included
short-term chronic toxicity tests for the fathead minnow,
Pimephales promelas, and the daphnid, Ceriodaphnia dubia. Birge et
al. (1985) reported that LC1s derived from Probit Analysis of data
from short-term embryo-larval tests with reference toxicants were
comparable to NOECs for several organisms. Similarly, USEPA (1988d)
reported that the IC25s were comparable to the NOECs for a set of
daphnid, Ceriodaphnia dubia, chronic tests with a single reference
toxicant. However, the scope of these comparisons was very limited,
and sufficient information is not yet available to establish an
overall relationship between these two types of endpoints,
especially when derived from effluent toxicity test
data.
9.3
PRECISION
9.3.1 HYPOTHESIS TESTS
9.3.1.1
When hypothesis tests are used to analyze toxicity test
data, it is not possible to express precision in terms of a
commonly used statistic. The results of the test are given in terms
of two endpoints, the No-Observed- Effect Concentration (NOEC) and
the Lowest-Observed-Effect Concentration (LOEC). The NOEC and LOEC
are limited to the concentrations selected for the test. The width
of the NOEC-LOEC interval is a function of the dilution series, and
differs greatly depending on whether a dilution factor of 0.3 or
0.5 is used in the test design. Therefore, USEPA recommends the use
of the $ 0.5 dilution factor (see Section 4, Quality Assurance). It
is not possible to place confidence limits on the NOEC and LOEC
derived from a given test, and it is difficult to quantify the
precision of the NOEC-LOEC endpoints between tests. If the data
from a series of tests performed with the same toxicant, toxicant
concentrations, and test species, were analyzed with hypothesis
tests, precision could only be assessed by a qualitative comparison
of the NOEC-LOEC intervals, with the understanding that maximum
precision would be attained if all tests yielded the same NOEC-LOEC
interval. In practice, the precision of results of repetitive
chronic tests is considered acceptable if the NOECs vary by no more
than one concentration interval above or below a central tendency.
Using these guidelines, the "normal" range of NOECs from toxicity
tests using a 0.5 dilution factor (two-fold difference between
adjacent concentrations), would be four-fold.
9.3.2
POINT ESTIMATION TECHNIQUES
9.3.2.1 Point estimation techniques have the advantage of
providing a point estimate of the toxicant concentration causing a
given amount of adverse (inhibiting) effect, the precision of which
can be quantitatively assessed
(1) within tests by calculation of 95% confidence limits, and
(2) across tests by calculating a standard deviation and
coefficient of variation.
9.3.2.2
It should be noted that software used to calculate point
estimates occasionally may not provide associated 95% confidence
intervals. This situation may arise when test data do not meet
specific assumptions required by the statistical methods, when
point estimates are outside of the test concentration range, and
when specific limitations imposed by the software are encountered.
USEPA (2000a) provides guidance on confidence intervals under these
circumstances.
9.4
DATA ANALYSIS
9.4.1 ROLE OF THE STATISTICIAN
9.4.1.1
The use of the statistical methods described in this
manual for routine data analysis does not require the assistance of
a statistician. However, the interpretation of the results of the
analysis of the data from any of the toxicity tests described in
this manual can become problematic because of the inherent
variability and sometimes unavoidable anomalies in biological data.
If the data appear unusual in any way, or fail to meet the
necessary assumptions, a statistician should be consulted. Analysts
who are not proficient in statistics are strongly advised to seek
the assistance of a statistician before selecting the method of
analysis and using any of the results.
9.4.1.2
The statistical methods recommended in this manual are
not the only possible methods of statistical analysis. Many other
methods have been proposed and considered. Certainly there are
other reasonable and defensible methods of statistical analysis for
this kind of toxicity data. Among alternative hypothesis tests
some, like Williams' Test, require additional assumptions, while
others, like the bootstrap methods, require computerintensive
computations. Alternative point estimation approaches most probably
would require the services of a statistician to determine the
appropriateness of the model (goodness of fit), higher order linear
or nonlinear models, confidence intervals for estimates generated
by inverse regression, etc. In addition, point estimation or
regression approaches would require the specification by biologists
or toxicologists of some low level of adverse effect that would be
deemed acceptable or safe. The statistical methods contained in
this manual have been chosen because they are (1) applicable to
most of the different toxicity test data sets for which they are
recommended, (2) powerful statistical tests, (3) hopefully "easily"
understood by nonstatisticians, and (4) amenable to use without a
computer, if necessary.
9.4.2
PLOTTING THE DATA
9.4.2.1
The data should be plotted, both as a preliminary step to
help detect problems and unsuspected trends or patterns in the
responses, and as an aid in interpretation of the results. Further
discussion and plotted sets of data are included in the methods and
the Appendices.
9.4.3
DATA TRANSFORMATIONS
9.4.3.1
Transformations of the data, (e.g., arc sine square root
and logs), are used where necessary to meet assumptions of the
proposed analyses, such as the requirement for normally distributed
data.
9.4.4
INDEPENDENCE, RANDOMIZATION, AND OUTLIERS
9.4.4.1
Statistical independence among observations is a critical
assumption in all statistical analysis of toxicity data. One of the
best ways to insure independence is to properly follow rigorous
randomization procedures. Randomization techniques should be
employed at the start of the test, including the randomization of
the placement of test organisms in the test chambers and
randomization of the test chamber location within the array of
chambers. Discussions of statistical independence, outliers and
randomization, and a sample randomization scheme, are included in
Appendix A.
9.4.5
REPLICATION AND SENSITIVITY
9.4.5.1
The number of replicates employed for each toxicant
concentration is an important factor in determining the sensitivity
of chronic toxicity tests. Test sensitivity generally increases as
the number of replicates is increased, but the point of diminishing
returns in sensitivity may be reached rather quickly. The level of
sensitivity required by a hypothesis test or the confidence
interval for a point estimate will determine the number of
replicates, and should be based on the objectives for obtaining the
toxicity data.
9.4.5.2
In a statistical analysis of toxicity data, the choice of
a particular analysis and the ability to detect departures from the
assumptions of the analysis, such as the normal distribution of the
data and homogeneity of variance, is also dependent on the number
of replicates. More than the minimum number of replicates may be
required in situations where it is imperative to obtain optimal
statistical results, such as with tests used in enforcement cases
or when it is not possible to repeat the tests. For example, when
the data are analyzed by hypothesis testing, the nonparametric
alternatives cannot be used unless there are at least four
replicates at each toxicant concentration.
9.4.6
RECOMMENDED ALPHA LEVELS
9.4.6.1 The data analysis examples included in the manual
specify an alpha level of 0.01 for testing the assumptions of
hypothesis tests and an alpha level of 0.05 for the hypothesis
tests themselves. These levels are common and well accepted levels
for this type of analysis and are presented as a recommended
minimum significance level for toxicity test data analysis.
9.5 CHOICE OF ANALYSIS
9.5.1
The recommended statistical analysis of most data from
chronic toxicity tests with aquatic organisms follows a decision
process illustrated in the flowchart in Figure 2. An initial
decision is made to use point estimation techniques (the Probit
Analysis, the Spearman-Karber Method, the Trimmed Spearman-Karber
Method, the Graphical Method, or Linear Interpolation Method)
and/or to use hypothesis testing (Dunnett's Test, the t test with
the Bonferroni adjustment, Steel's Many-one Rank Test, or the
Wilcoxon Rank Sum Test with the Bonferroni adjustment). NOTE: For
the NPDES Permit Program, the point estimation techniques are the
preferred statistical methods in calculating end points for
effluent toxicity tests. If hypothesis testing is chosen,
subsequent decisions are made on the appropriate procedure for a
given set of data, depending on the results of the tests of
assumptions, as illustrated in the flowchart. A specific flow chart
is included in the analysis section for each test.
9.5.2
Since a single chronic toxicity test might yield
information on more than one parameter (such as survival, growth,
and reproduction), the lowest estimate of a "no-observed-effect
concentration" for any of the responses would be used as the
"no-observed-effect concentration" for each test. It follows
logically that in the statistical analysis of the data,
concentrations that had a significant toxic effect on one of the
observed responses would not be subsequently tested for an effect
on some other response. This is one reason for excluding
concentrations that have shown a statistically significant
reduction in survival from a subsequent hypothesis test for effects
on another parameter such as reproduction. A second reason is that
the exclusion of such concentrations usually results in a more
powerful and appropriate statistical analysis. In performing the
point estimation techniques recommended in this manual, an all-data
approach is used. For example, data from concentrations above the
NOEC for survival are included in determining ICp estimates using
the Linear Interpolation Method.
9.5.3
ANALYSIS OF GROWTH AND REPRODUCTION DATA
9.5.3.1
Growth data from the fathead minnow, Pimephales promelas,
larval survival and growth test are analyzed using hypothesis
testing or point estimation techniques according to the flowchart
in Figure 2. The above mentioned growth data may also be analyzed
by generating a point estimate with the Linear Interpolation
Method. Data from effluent concentrations that have tested
significantly different from the control for survival are excluded
from further hypothesis tests concerning growth effects. Growth is
defined as the dry weight per original number of test organisms
when group weights are obtained. When analyzing the data using
point estimation techniques, data from all concentrations are
included in the analysis.
9.5.3.2
Reproduction data from the daphnid, Ceriodaphnia dubia,
survival and reproduction test are analyzed using hypothesis
testing or point estimation techniques according to the flowchart
in Figure 2. In hypothesis testing, data from effluent
concentrations that have significantly lower survival than the
control, as determined by Fisher's Exact test, are not included in
the hypothesis tests for reproductive effects. Data from all
concentrations are included when using point estimation
techniques.
9.5.4
ANALYSIS OF ALGAL GROWTH RESPONSE DATA
9.5.4.1
The growth response data from the green alga, Selenastrum
capricornutum, toxicity test, after an appropriate transformation,
if necessary, to meet the assumptions of normality and homogeneity
of variance, may be analyzed by hypothesis testing according to the
flowchart in Figure 2. Point estimates, such as the IC25 and IC50,
would also be appropriate in analyzing algal growth
data.
9.5.5
ANALYSIS OF MORTALITY DATA
9.5.5.1
Mortality data are analyzed by Probit Analysis, if
appropriate, or other point estimation techniques (i.e., the
Spearman-Karber Method, the Trimmed Spearman-Karber Method, or the
Graphical Method) (see Appendices I-L and the discussion below).
The mortality data can also be analyzed by hypothesis testing,
after an arc sine square root transformation (see Appendix B-F),
according to the flowchart in Figure 2.
9.5.5.2
Mortality data from the daphnid, Ceriodaphnia dubia,
survival and reproduction test are analyzed by Fisher's Exact Test
(Appendix G) prior to the analysis of the reproduction data. The
mortality data may also be analyzed by Probit Analysis, if
appropriate or other methods (see Subsection 9.5.5.1).
9.6
HYPOTHESIS TESTS
9.6.1 DUNNETT'S PROCEDURE
9.6.1.1
Dunnett's Procedure is used to determine the NOEC. The
procedure consists of an analysis of variance (ANOVA) to determine
the error term, which is then used in a multiple comparison
procedure for comparing each of the treatment means with the
control mean, in a series of paired tests (see Appendix C). Use of
Dunnett's Procedure requires at least three replicates per
treatment to check the assumptions of the test. In cases where the
numbers of data points (replicates) for each concentration are not
equal, a t test may be performed with Bonferroni's adjustment for
multiple comparisons (see Appendix D), instead of using Dunnett's
Procedure.
9.6.1.2
The assumptions upon which the use of Dunnett's Procedure
is contingent are that the observations within treatments are
normally distributed, with homogeneity of variance. Before
analyzing the data, these assumptions must be tested using the
procedures provided in Appendix B.
9.6.1.3
If, after suitable transformations have been carried out,
the normality assumptions have not been met, Steel's Many-one Rank
Test should be used if there are four or more data points
(replicates) per toxicant concentration. If the numbers of data
points for each toxicant concentration are not equal, the Wilcoxon
Rank Sum Test with Bonferroni's adjustment should be used (see
Appendix F).
9.6.1.4
Some indication of the sensitivity of the analysis should
be provided by calculating (1) the minimum difference between means
that can be detected as statistically significant, and (2) the
percent change from the control mean that this minimum difference
represents for a given test.
9.6.1.5
A step-by-step example of the use of Dunnett's Procedure
is provided in Appendix C.
9.6.2
T TEST WITH THE BONFERRONI ADJUSTMENT
9.6.2.1
A t test with Bonferroni's adjustment is used as an
alternative to Dunnett's Procedure when the number of replicates is
not the same for all concentrations. This test sets an upper bound
of alpha on the overall error rate, in contrast to Dunnett's
Procedure, for which the overall error rate is fixed at alpha. Thus
Dunnett's Procedure is a more powerful test.
9.6.2.2
The assumptions upon which the use of the t test with
Bonferroni's adjustment is contingent are that the observations
within treatments are normally distributed, with homogeneity of
variance. These assumptions must be tested using the procedures
provided in Appendix B.
9.6.2.3
The estimate of the safe concentration derived from this
test is reported in terms of the NOEC. A step-by-step example of
the use of the t test with Bonferroni's adjustment is provided in
Appendix D.
9.6.3
STEEL'S MANY-ONE RANK TEST
9.6.3.1
Steel's Many-one Rank Test is a multiple comparison
procedure for comparing several treatments with a control. This
method is similar to Dunnett's Procedure, except that it is not
necessary to meet the assumption of normality. The data are ranked,
and the analysis is performed on the ranks rather than on the data
themselves. If the data are normally or nearly normally
distributed, Dunnett's Procedure would be more sensitive (would
detect smaller differences between the treatments and control). For
data that are not normally distributed, Steel's Many-one Rank Test
can be much more efficient (Hodges and Lehmann, 1956).
9.6.3.2
It is necessary to have at least four replicates per
toxicant concentration to use Steel's test. Unlike Dunnett's
procedure, the sensitivity of this test cannot be stated in terms
of the minimum difference between treatment means and the control
mean that can be detected as statistically significant.
9.6.3.3
The estimate of the safe concentration is reported as the
NOEC. A step-by-step example of the use of Steel's Many-one Rank
Test is provided in Appendix E.
9.6.4
WILCOXON RANK SUM TEST WITH THE BONFERRONI
ADJUSTMENT
9.6.4.1
The Wilcoxon Rank Sum Test with the Bonferroni Adjustment
is a nonparametric test for comparing treatments with a control.
The data are ranked and the analysis proceeds exactly as in Steel's
Test except that Bonferroni's adjustment for multiple comparisons
is used instead of Steel's tables. When Steel's test can be used
(i.e., when there are equal numbers of data points per toxicant
concentration), it will be more powerful (able to detect smaller
differences as statistically significant) than the Wilcoxon Rank
Sum Test with Bonferroni's adjustment.
9.6.4.2
The estimate of the safe concentration is reported as the
NOEC. A step-by-step example of the use of the Wilcoxon Rank Sum
Test with Bonferroni Adjustment is provided in Appendix
F.
9.6.5
A CAUTION IN THE USE OF HYPOTHESIS TESTING
9.6.5.1
If in the calculation of an NOEC by
hypothesis testing, two tested concentrations cause statistically
significant adverse effects, but an intermediate concentration did
not cause statistically significant effects, the results should be
used with extreme caution.
9.7
POINT ESTIMATION TECHNIQUES
9.7.1 PROBIT ANALYSIS
9.7.1.1
Probit Analysis is used to estimate the LC1, LC50, EC1,
or EC50 and the associated 95% confidence interval. The analysis
consists of adjusting the data for mortality in the control, and
then using a maximum likelihood technique to estimate the
parameters of the underlying log tolerance distribution, which is
assumed to have a particular shape.
9.7.1.2
The assumption upon which the use of Probit Analysis is
contingent is a normal distribution of log tolerances. If the
normality assumption is not met, and at least two partial
mortalities are not obtained, Probit Analysis should not be used.
It is important to check the results of Probit Analysis to
determine if use of the analysis is appropriate. The chi-square
test for heterogeneity provides one good test of appropriateness of
the analysis. The computer program (see Appendix I) checks the
chi-square statistic calculated for the data set against the
tabular value, and provides an error message if the calculated
value exceeds the tabular value.
9.7.1.3
A discussion of Probit Analysis, and examples of computer
program input and output, are found in Appendix I.
9.7.1.4
In cases where Probit Analysis is not appropriate, the
LC50 and associated confidence interval may be estimated by the
Spearman-Karber Method (Appendix J) or the Trimmed Spearman-Karber
Method (Appendix K). If the test results in 100% survival and 100%
mortality in adjacent treatments (all or nothing effect), the LC50
may be estimated using the Graphical Method (Appendix
L).
9.7.2
LINEAR INTERPOLATION METHOD
9.7.2.1
The Linear Interpolation Method (see Appendix M) is a
procedure to calculate a point estimate of the effluent or other
toxicant concentration [Inhibition Concentration, (IC)] that causes
a given percent reduction (e.g., 25%, 50%, etc.) in the
reproduction or growth of the test organisms. The procedure was
designed for general applicability in the analysis of data from
short-term chronic toxicity tests.
9.7.2.2
Use of the Linear Interpolation Method is based on the
assumptions that the responses (1) are monotonically non-increasing
(the mean response for each higher concentration is less than or
equal to the mean response for the previous concentration), (2)
follow a piecewise linear response function, and (3) are from a
random, independent, and representative sample of test data. The
assumption for piecewise linear response cannot be tested
statistically, and no defined statistical procedure is provided to
test the assumption for monotonicity. Where the observed means are
not strictly monotonic by examination, they are adjusted by
smoothing. In cases where the responses at the low toxicant
concentrations are much higher than in the controls, the smoothing
process may result in a large upward adjustment in the control
mean.
9.7.2.3
The inability to test the monotonicity and piecewise
linear response assumptions for this method makes it difficult to
assess when the method is, or is not, producing reliable results.
Therefore, the method should be used with caution when the results
of a toxicity test approach an "all or nothing" response from one
concentration to the next in the concentration series, and when it
appears that there is a large deviation from monotonicity. See
Appendix M for a more detailed discussion of the use of this method
and a computer program available for performing
calculations.
SECTION 10
REPORT PREPARATION AND TEST REVIEW
10.1 REPORT PREPARATION
The following general format and content are recommended for the
report:
10.1.1 INTRODUCTION
1.
Permit number
2.
Toxicity testing requirements of permit
3.
Plant location
4.
Name of receiving water body
5.
Contract Laboratory (if the tests are performed under
contract) a Name of firm
b.
Phone number
c.
Address
6. Objective of test
10.1.2 PLANT OPERATIONS
1.
Product(s)
2.
Raw materials
3.
Operating schedule
4.
Description of waste treatment
5.
Schematic of waste treatment
6.
Retention time (if applicable)
7.
Volume of waste flow (MGD, CFS, GPM)
8.
Design flow of treatment facility at time of
sampling
10.1.3 SOURCE OF EFFLUENT, RECEIVING WATER, AND DILUTION
WATER
1. Effluent Samples
a.
Sampling point (including latitude and
longitude)
b.
Collection dates and times
c.
Sample collection method
d.
Physical and chemical data
e.
Mean daily discharge on sample collection date
f.
Lapsed time from sample collection to delivery
g.
Sample temperature when received at the
laboratory
2. Receiving Water Samples
a.
Sampling point (including latitude and
longitude)
b.
Collection dates and times
c.
Sample collection method
d.
Physical and chemical data
e.
Streamflow (at time of sampling)
f.
Sample temperature when received at the laboratory g
Lapsed time from sample collection to delivery
3. Dilution Water Samples
a.
Source
b.
Collection date(s) and time(s)
c.
Pretreatment
d.
Physical and chemical characteristics
10.1.4 TEST METHODS
1.
Toxicity test method used (title, number,
source)
2.
Endpoint(s) of test
3.
Deviation(s) from reference method, if any, and the
reason(s)
4.
Date and time test started
5.
Date and time test terminated
6.
Type and volume of test chambers
7.
Volume of solution used per chamber
8.
Number of organisms per test chamber
9.
Number of replicate test chambers per
treatment
10.
Acclimation of test organisms (temperature mean and
range)
11.
Test temperature (mean and range)
12.
Specify if aeration was needed
13.
Feeding frequency, and amount and type of food
14.
Specify if (and how) pH control measures were
implemented
10.1.5 TEST ORGANISMS
1.
Scientific name and how determined
2.
Age
3.
Life stage
4.
Mean length and weight (where applicable)
5.
Source
6.
Diseases and treatment (where applicable)
7.
Taxonomic key used for species identification
10.1.6 QUALITY ASSURANCE
1.
Reference toxicant used routinely; source
2.
Date and time of most recent reference toxicant test,
test results, and current control chart
3.
Dilution water used in reference toxicant test
4.
Results (NOEC or, where applicable, LOEC, LC50, EC50,
IC25 and/or IC50); report percent minimum significant difference
(PMSD) calculated for sublethal endpoints determined by hypothesis
testing in reference toxicant test
5.
Physical and chemical methods used
10.1.7 RESULTS
1.
Provide raw toxicity data in tabular form, including
daily records of affected organisms in each concentration
(including controls) and replicate, and in graphical form (plots of
toxicity data)
2.
Provide table of LC50s, NOECs, IC25, IC50, etc. (as
required in the applicable NPDES permit)
3.
Indicate statistical methods used to calculate
endpoints
4.
Provide summary table of physical and chemical
data
5.
Tabulate QA data
6.
Provide percent minimum significant difference (PMSD)
calculated for sublethal endpoints
10.1.8 CONCLUSIONS AND RECOMMENDATIONS
1.
Relationship between test endpoints and permit
limits
2.
Actions to be taken
10.2 TEST REVIEW
10.2.1
Test review is an important part of an overall quality
assurance program (Section 4) and is necessary for ensuring that
all test results are reported accurately. Test review should be
conducted on each test by both the testing laboratory and the
regulatory authority.
10.2.2
SAMPLING AND HANDLING
10.2.2.1
The collection and handling of samples are reviewed to
verify that the sampling and handling procedures given in Section 8
were followed. Chain-of-custody forms are reviewed to verify that
samples were tested within allowable sample holding times
(Subsection 8.5.4). Any deviations from the procedures given in
Section 8 should be documented and described in the data report
(Subsection 10.1).
10.2.3
TEST ACCEPTABILITY CRITERIA
10.2.3.1
Test data are reviewed to verify that test acceptability
criteria (TAC) requirements for a valid test have been met. Any
test not meeting the minimum test acceptability criteria is
considered invalid. All invalid tests must be repeated with a newly
collected sample.
10.2.4
TEST CONDITIONS
10.2.4.1
Test conditions are reviewed and compared to the
specifications listed in the summary of test condition tables
provided for each method. Physical and chemical measurements taken
during the test (e.g., temperature, pH, and DO) also are reviewed
and compared to specified ranges. Any deviations from
specifications should be documented and described in the data
report (Subsection 10.1).
10.2.4.2
The summary of test condition tables presented for each
method identify test conditions as required or recommended. For WET
test data submitted under NPDES permits, all required test
conditions must be met or the test is considered invalid and must
be repeated with a newly collected sample. Deviations from
recommended test conditions must be evaluated on a case-by-case
basis to determine the validity of test results. Deviations from
recommended test conditions may or may not invalidate a test result
depending on the degree of the departure and the objective of the
test. The reviewer should consider the degree of the deviation and
the potential or observed impact of the deviation on the test
result before rejecting or accepting a test result as valid. For
example, if dissolved oxygen is measured below 4.0 mg/L in one test
chamber, the reviewer should consider whether any observed
mortality in that test chamber corresponded with the drop in
dissolved oxygen.
10.2.4.3
Whereas slight deviations in test conditions may not
invalidate an individual test result, test condition deviations
that continue to occur frequently in a given laboratory may
indicate the need for improved quality control in that
laboratory.
10.2.5
STATISTICAL METHODS
10.2.5.1 The statistical methods used for analyzing test data
are reviewed to verify that the recommended flowcharts for
statistical analysis were followed. Any deviation from the
recommended flowcharts for selection of statistical methods should
be noted in the data report. Statistical methods other than those
recommended in the statistical flowcharts may be appropriate (see
Subsection 9.4.1.2), however, the laboratory must document the use
of and provide the rationale for the use of any alternate
statistical method. In all cases (flowchart recommended methods or
alternate methods), reviewers should verify that the necessary
assumptions are met for the statistical method used.
10.2.6 CONCENTRATION-RESPONSE RELATIONSHIPS
10.2.6.1
The concept of a concentration-response, or more
classically, a dose-response relationship is "the most fundamental
and pervasive one in toxicology" (Casarett and Doull, 1975). This
concept assumes that there is a causal relationship between the
dose of a toxicant (or concentration for toxicants in solution) and
a measured response. A response may be any measurable biochemical
or biological parameter that is correlated with exposure to the
toxicant. The classical concentration-response relationship is
depicted as a sigmoidal shaped curve, however, the particular shape
of the concentration-response curve may differ for each coupled
toxicant and response pair. In general, more severe responses (such
as acute effects) occur at higher concentrations of the toxicant,
and less severe responses (such as chronic effects) occur at lower
concentrations. A single toxicant also may produce multiple
responses, each characterized by a concentration-response
relationship. A corollary of the concentration-response concept is
that every toxicant should exhibit a concentration-response
relationship, given that the appropriate response is measured and
given that the concentration range evaluated is appropriate. Use of
this concept can be helpful in determining whether an effluent
possesses toxicity and in identifying anomalous test
results.
10.2.6.2
The concentration-response relationship generated for
each multi-concentration test must be reviewed to ensure that
calculated test results are interpreted appropriately. USEPA
(2000a) provides guidance on evaluating concentration-response
relationships to assist in determining the validity of WET test
results. All WET test results (from multi-concentration tests)
reported under the NPDES program should be reviewed and reported
according to USEPA guidance on the evaluation of
concentration-response relationships (USEPA, 2000a). This guidance
provides review steps for 10 different concentration-response
patterns that may be encountered in WET test data. Based on the
review, the guidance provides one of three determinations: that
calculated effect concentrations are reliable and should be
reported, that calculated effect concentrations are anomalous and
should be explained, or that the test was inconclusive and the test
should be repeated with a newly collected sample. It should be
noted that the determination of a valid concentration-response
relationship is not always clear cut. Data from some tests may
suggest consultation with professional toxicologists and/or
regulatory officials. Tests that exhibit unexpected
concentration-response relationships also may indicate a need for
further investigation and possible retesting.
10.2.7
REFERENCE TOXICANT TESTING
10.2.7.1
Test review of a given effluent or receiving water test
should include review of the associated reference toxicant test and
current control chart. Reference toxicant testing and control
charting is required for documenting the quality of test organisms
(Subsection 4.7) and ongoing laboratory performance (Subsection
4.16). The reviewer should verify that a quality control reference
toxicant test was conducted according to the specified frequency
required by the permitting authority or recommended by the method
(e.g., monthly). The test acceptability criteria, test conditions,
concentration-response relationship, and test sensitivity of the
reference toxicant test are reviewed to verify that the reference
toxicant test conducted was a valid test. The results of the
reference toxicant test are then plotted on a control chart (see
Subsection 4.16) and compared to the current control chart limits
(± 2 standard deviations).
10.2.7.2
Reference toxicant tests that fall outside of recommended
control chart limits are evaluated to determine the validity of
associated effluent and receiving water tests (see Subsection
4.16). An out of control reference toxicant test result does not
necessarily invalidate associated test results. The reviewer should
consider the degree to which the reference toxicant test result
fell outside of control chart limits, the width of the limits, the
direction of the deviation (toward increasing test organism
sensitivity or toward decreasing test organism sensitivity), the
test conditions of both the effluent test and the reference
toxicant test, and the objective of the test. More frequent and/or
concurrent reference toxicant testing may be advantageous if recent
problems (e.g., invalid tests, reference toxicant test results
outside of control chart limits, reduced health of organism
cultures, or increased within-test variability) have been
identified in testing.
10.2.8
TEST VARIABILITY
10.2.8.1
The within-test variability of individual tests should be
reviewed. Excessive within-test variability may invalidate a test
result and warrant retesting. For evaluating within-test
variability, reviewers should consult EPA guidance on upper and
lower percent minimum significant difference (PMSD) bounds (USEPA,
2000b).
10.2.8.2
When NPDES permits require sublethal hypothesis testing
endpoints from Methods 1000.0,1002.0, or 1003.0 (e.g., growth or
reproduction NOECs and LOECs), within-test variability must be
reviewed and variability criteria must be applied as described in
this section (10.2.8.2). When the methods are used for
non-regulatory purposes, the variability criteria herein are
recommended but are not required, and their use (or the use of
alternative variability criteria) may depend upon the intended uses
of the test results and the requirements of any applicable data
quality objectives and quality assurance plan.
10.2.8.2.1
To measure test variability, calculate the percent
minimum significant difference (PMSD) achieved in the test. The
PMSD is the smallest percentage decrease in growth or reproduction
from the control that could be determined as statistically
significant in the test. The PMSD is calculated as 100 times the
minimum significant difference (MSD) divided by the control mean.
The equation and examples of MSD calculations are shown in Appendix
C. PMSD may be calculated legitimately as a descriptive statistic
for within-test variability, even when the hypothesis test is
conducted using a non-parametric method. The PMSD bounds were based
on a representative set of tests, including tests for which a
non-parametric method was required for determining the NOEC or
LOEC. The conduct of hypothesis testing to determine test results
should follow the statistical flow charts provided for each method.
That is, when test data fail to meet assumptions of normality or
heterogeneity of variance, a nonparametric method (determined
following the statistical flowchart for the method) should be used
to calculate test results, but the PMSD may be calculated as
described above (using parametric methods) to provide a measure of
test variability.
10.2.8.2.2
Compare the PMSD measured in the test with the upper PMSD
bound variability criterion listed in Table 6. When the test PMSD
exceeds the upper bound, the variability among replicates is
unusually large for the test method. Such a test should be
considered insufficiently sensitive to detect toxic effects on
growth or reproduction of substantial magnitude. A finding of
toxicity at a particular concentration may be regarded as
trustworthy, but a finding of "no toxicity" or "no statistically
significant toxicity" at a particular concentration should not be
regarded as a reliable indication that there is no substantial
toxic effect on growth or reproduction at that
concentration.
10.2.8.2.3
If the PMSD measured for the test is less than or equal
to the upper PMSD bound variability criterion in Table 6, then the
test's variability measure lies within normal bounds and the effect
concentration estimate (e.g., NOEC or LOEC) would normally be
accepted unless other test review steps raise serious doubts about
its validity.
10.2.8.2.4
If the PMSD measured for the test exceeds the upper PMSD
bound variability criterion in Table 6, then one of the following
two cases applies (10.2.8.2.4.1, 10.2.8.2.4.2).
10.2.8.2.4.1
If toxicity is found at the permitted receiving water
concentration (RWC) based upon the value of the effect
concentration estimate (NOEC or LOEC), then the test shall be
accepted and the effect concentration estimate may be reported,
unless other test review steps raise serious doubts about its
validity.
10.2.8.2.4.2
If toxicity is not found at the permitted RWC based upon
the value of the effect concentration estimate (NOEC or LOEC) and
the PMSD measured for the test exceeds the upper PMSD bound, then
the test shall not be accepted, and a new test must be conducted
promptly on a newly collected sample.
10.2.8.2.5
To avoid penalizing laboratories that achieve unusually
high precision, lower PMSD bounds shall also be applied when a
hypothesis test result (e.g., NOEC or LOEC) is reported. Lower PMSD
bounds, which are based on the 10th percentiles of national PMSD
data, are presented in Table 6. The 10th percentile PMSD represents
a practical limit to the sensitivity of the test method because few
laboratories are able to achieve such precision on a
regular basis and most do not achieve it even occasionally. In
determining hypothesis test results (e.g., NOEC or LOEC), a test
concentration shall not be considered toxic (i.e., significantly
different from the control) if the relative difference from the
control is less than the lower PMSD bounds in Table 6. See USEPA,
2000b for specific examples of implementing lower PMSD bounds.
10.2.8.3 To assist in reviewing within-test variability, EPA
recommends maintaining control charts of PMSDs calculated for
successive effluent tests (USEPA, 2000b). A control chart of PMSD
values characterizes the range of variability observed within a
given laboratory, and allows comparison of individual test PMSDs
with the laboratory's typical range of variability. Control charts
of other variability and test performance measures, such as the
MSD, standard deviation or CV of control responses, or average
control response, also may be useful for reviewing tests and
minimizing variability. The log of PMSD will provide an
approximately normal variate useful for control charting.
TABLE 6. VARIABILITY CRITERIA (UPPER AND LOWER PMSD BOUNDS) FOR
SUBLETHAL HYPOTHESIS TESTING ENDPOINTS SUBMITTED UNDER NPDES
PERMITS.1
1 Lower and upper PMSD bounds were determined from the 10th and
90th percentile, respectively, of PMSD data from EPA's WET
Interlaboratory Variability Study (USEPA, 2001a; USEPA, 2001b).
