SEPA

IJnlted States Office of Research and

Enwronmental Protection Development

Agency Washington DC 20460

EPA,‘600/R-92/08 1

September 1993

Methods for Aquatic

Toxicity Identification

Evaluations

Phase Ill Toxicity

Confirmation Procedures for

Samples Exhibiting Acute and

Chronic Toxicity c
EPA/600/R-92/081

September 1993

Methods for Aquatic

Toxicity Identification Evaluations

Phase I I I Toxicity Confirmation

Procedures for Samples Exhibiting

Acute and Chronic Toxicity

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bY

D. I. Mount’

T. J. Norberg-King*

With Contributions from:

G. T. Ankley*

L. P. Burkhard2

E. J. Durban*

M. K. Schubauer-Beriganl

M. T. Lukasewyczl

‘AScl Corporation - Contract No. 68-CO-0058

*U.S. Environmental Protection Agency

Prev’iis Phase Ill Methods

by 0. I. Mount

EPA-600/3-88/036

National Effluent Toxicii Assessment Center

Technical Report 02-93

Environmental Research Laboratory

Office of Research and Development

U.S. Environmental Protection Agency

Duluth, MN 55804

@ Printed on Recycled Paper
Disclaimer

This document has been reviewed in accordance with U.S. Environmental

Protection Agency Policy and approved for publication. Mention of trade names or

commercial products does not constitute endorsement or recommendation for use.
Foreword

This Phase III document is the last in a series of guidance documents

intended to aid dischargers and their consultants in conducting aquatic organism

toxicity identification evaluations (TIES). TIES might be required by state or federal

agencies as the result of an enforcement action or as a condition of a National

Pollutant Discharge Elimination System (NPDES) permit. These documents should

aid individuals in overseeing and determining the adequacy of effluent TIES as a part

of toxicity reduction evaluations (TREs).

There are two major reasons to require the confirmation procedures. First the

effluent manipulations used in Phase I characterizations (EPA, 1988; EPA, 1991 A;

EPA, 1992) and Phase II identifications (EPA, 1989A; EPA, 1993A) might (with some

effluents) create artifacts that might lead to erroneous conclusions about the cause

of toxicity. Therefore in Phase III confirmation steps, manipulations of the effluent are

avoided and/or are minimized, therefore artifacts are far less likely to occur. Sometimes, toxicants will be suspected through other approaches (such as the treatability

route) which on their own are not definitive and in these instances, confirmation is

necessary. Secondly, there is the probability that the substances causing toxicity

might change from sample to sample, from season to season or some other

periodicity. As toxicity is a generic measurement, measuring toxicity cannot reveal

variability of the suspect toxicant whereas the Phase III confirmation procedures are

designed to indicate the presence of variable toxicants. Obviously, this crucial

information is essential so that remedial action may be taken to remove toxicity.

t

Confirmation, whether using the procedures described in this document or

others, should always be completed because the risk is too great to avoid or eliminate

this step. Especially for discharges where there is little control over the influent or for

discharge operations that are very large or complex, the probability that different

constituents tiill cause toxicity over time is great. Most of the approaches in Phase III

are applicable to chronically toxic effluents and acutely toxic effluents.

In this confirmation document, guidance is included when the treatability

approach (EPA, 19898; EPA, 1989C) is taken. Use of the treatability approach

requires confirmation as much as or more than the toxicant identification approach

(Phase II). The reader is encouraged to use both the acute Phase I characterization

(EPA, 1991 A) and the chronic Phase I characterization (EPA, 1992) documents for

details of quality assurance/quality control (CWQC), health and safety, facilities and

equipment, dilution water, sampling and testing. The TIE methods are written as

general guidance rather than rigid protocols for conducting TIES and these methods

should be applicable to other aqueous samples, such as ambient waters, sediment

” elutriate or pore waters, and 1eachate.s

. . .

III
Abstract

In 1989, the guidance document for acutely toxic effluents entitled Methods

for Aquatic Toxicity Identification Evaluations: Phase Ill Toxicity Confirmation Procedures was published (EPA, 19890). This new Phase Ill manual and its companion

documents (EPA, 1991 A; EPA, 1992; EPA, 1993A) are intended to provide guidance

to aid dischargers in confirming the cause of toxicity in industrial and municipal

effluents. The toxicity identification evaluation (TIE) starts with a characterization of

the effluent toxicity using aquatic organisms to track toxicity; this step is followed by

identifying a suspect toxicant and then confirming the suspect toxicant as the cause

of toxicity.

This Phase III confirmation document provides greater detail and more

insight into the procedures described in the acute Phase III confirmation document

(EPA, 1989D). Procedures to confirm that all toxicants have been correctly identified

are given and specific changes for methods applicable to chronic toxicity are included.

Adifficult aspect of confirmation occurs when toxicants are not additive, and therefore

the effects of effluent matrix affecting the toxicants are discussed. The same basic

techniques (correlation, symptoms, relative species sensitivity, spiking, and mass

balance) are still used to confirm toxicants and case examples are provided to

illustrate some of the Phase III procedures. Procedures that describe the techniques

to characterize the acute or chronic toxicity (EPA, 1988) and to identify (EPA, 1989A)

toxicants have also been rewritten (EPA, 1991 A; EPA, 1992; EPA, 1993A).

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iv

c
Contents

Page

Foreword ....................................................................................................................... iii

Abstract ......................................................................................................................... iv

Tables ............................................................................................................................ vi

Figures ......................................................................................................................... .vi

Acknowledgments ....................................................................................................... .vii

1.

2.

3.

4.

5.

6.

7.

8.

9.

10.

11.

12.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1-l

Correlation Approach ......................................................................................... 2-l

2.1 Correlation ................................................................................................ ..2- 1

2.2 Correlation Problems Caused by Matrix Effects ........................................ .2-4

Symptom Approach.......................................................................................... ..3- 1

Species Sensitivity Approach ............................................................................ .4-l

Spiking Approach .............................................................................................. .5- 1

Mass Balance Approach . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-l

Deletion Approach.. ........................................................................................... .7-l

Additional Approaches ...................................................................................... 8-I

Hidden Toxicants .............................................................................................. .9-l

Conclusions ...................................................................................................... 1 o-1

When the Treatability Approach Has Been Used . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11-1

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12-1
,

,

Tables

Number Page

6-l. Comparison of effluent toxicity and toxicity measured in effluent

fraction add-back tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-2

L

Figures

Number

2-l.

2-2.

2-3.

2-4.

2-5.

Page

Correlation of toxic units (TUs) for an effluent and one suspect toxicant

in a POTW effluent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-2

Correlation of toxic units (TUs) for an effluent and one suspect toxicant

in a POTW effluent when two toxicants are the cause of toxicity . . . . . . . . . . . . . . . . . 2-2

Correlation of toxic units (TUs) for an effluent and two toxicants in a

POTW effluent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-3

Correct (top) and incorrect (bottom) plots of toxic units (TUs) for nonadditive toxicants . . . . . . . . . . . . . . . ..*........................................................................ 2-4

Correlation of toxic units (TUs) for a POTW effluent and the suspect

toxicant, nickel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-5

vi
6

Acknowledgments

This document presents additional information acquired since the document

entitled Methods for Aquatic Toxicity Identification Evaluations: Phase III Toxicity

Confirmation Procedures (EPA-60013-88-036; EPA, 1989D) was prepared by Donald

Mount and published in 1989. This manual reflects new information, techniques, and

suggestions made since the Phase III confirmation methods for acute toxicity were

developed. The suggestions, techniques and cautions contained in this document are

based on a large database generated by the staff of the National Effluent Toxicity

Assessment Center (NETAC) at the U.S. Environmental Protection Agency (EPA),

Environmental Research Laboratory, Duluth (ERL-D), MN. NETAC staff that provided

technical support consisted of Penny Juenemann and Shaneen Schmidt (ERL-D

staff), Joe Amato, Lara Anderson, Steve Baker, Tim Dawson, Nola Englehorn, Doug

Jensen, Correne Jenson, Jim Jenson, Elizabeth Makynen, Phil Monson, Greg

Peterson, and Jo Thompson (contract staff). Their collective experience has made

this document possible and the contributions are gratefully acknowledged. The

support through EPA’s Office of Research and Development (ORD) and Office of

Water made this research possible at ERL-D.

L

Vii
.

Section 1

Introduction

The final confirmation phase of a toxicity identification evaluation (TIE) consists of a group of steps

intended to confirm that the suspect cause(s) of toxicity is

correctly identified and that all the toxicity is accounted

for. Typically this confirmation step follows experiments

from the toxicity characterization step (Phase I) and

analysis and additional experiments conducted in toxicity

identification (Phase II) (EPA, 1991A; EPA, 1992; EPA,

1993A). However, there often may be no identifiable

boundary between phases. In fact, all three phases might

be underway concurrently with each effluent sample and

depending on the results of Phase I characterization, the

Phase II identification, and Phase III confirmation activities might begin with the first sample evaluated. Phase Ill

confirmation procedures should also follow after toxicants have been identified by other means or when

treatability approaches are used. Rarely does one step or

one test conclusively prove the cause of toxicity in Phase

III. Rather, all practical approaches are used to provide

the weight of evidence that the cause of toxicity has been

identified. The various approaches that are often useful

in providing that weight of evidence consist of correlation,

observation of symptoms, relative species sensitivity,

spiking, mass balance estimates and various adjustments of water quality.

The approaches described in this document have

been useful in TIES at ERL-D. While the guidance provided in this manual is based largely on experience with

wastewater effluents, in general the methods discussed

are applicable to ambient waters (Norberg-King et al.,

1991) and sediment pore or elutriate water samples as

well (EPA, 19918). However, specific modifications of

the TIE techniques might be needed (e.g., sample volume) when evaluating these other types of samples.

Confirmation is important to provide data to prove

that the suspect toxicant is the cause of toxicity in a

series of samples and to assure that all other toxicants

are identified that might occur in any sample over time.

There may be a tendency to assume that toxicity is

always caused by the same constituents, and if this

assumption carries over into the data interpretation but

the assumption is false, erroneous conclusions might be

reached. That is why the correlation step (Section 2) is

accompanied by other approaches (i.e., Sections 3-9)

because each approach aids in revealing any changes in

the toxicant in the confirmation phase of the TIE.

Seasonal trends in toxicants have been observed

in publicly owned treatment works (POTW) effluents and

some sediment samples. For example, organophosphate

pesticides have been observed to increase in concentrations in wastewaters during the late winter and spring

months (Norberg-King et al., 1989). Therefore, the confirmation steps of Phase III might need to include seasonal

samples. This effort cannot always be pre-determined.

The presence of a different toxicant must be considered throughout the TIE, and when samples are collected

over several months the seasonality of a suspect toxicant

should be carefully considered and studied. When remedial action requires treatment changes, one must be

certain that toxicity from specific toxicant is consistently

present and that the suspect toxicant accounts for all

the toxicity. Treatment modifications will not necessarily

result in removal of all toxicants to acceptable concentrations. If toxicity is caused by a variety of toxicants present

at varying intervals, the remedial actions that are practical

might differ from the remedial action required when toxicity is caused by the same constituents consistently.

c

TIES conducted at ERL-D have shown that toxicants often are not additive or toxicants are present in

ratios such that the toxicity contribution by one might be

diluted out in the range of the effluent effect concentration

(e.g., LC50 or ICp value). Thus, the toxicant present at

lower yet toxic concentrations may not be readily discerned. The frequency of occurrence and impact on data

interpretation of either of the above cases was not addressed previously (EPA, 1989D) but are now discussed

in Section 2. Toxicants that do not express their toxicity

because of the presence of other toxicants (either the

toxicants are non-additive or the toxicants occur in disparate ratios) are referred to as hidden toxicants (Section 9).

Detection of hidden toxicants is one of the most difficult

aspects of confirmation. It is a mistake to search for a

concentration of any chemical present in the effluent at a

toxic concentration and to declare any found as the cause

l-l
of toxicity. Matrix effects of the effluent samples make

conclusions such as these subject to error without further

work as either the hidden toxicant or the principal

toxicant are likely to be missed using such an approach.

There is a strong tendency to shorten or eliminate

the confirmation steps because by the time Phase III

confirmation has been reached, the investigators might

be convinced of the cause of toxicity and the confirmation

steps seem redundant. However, one cannot expect to

concentrate the effluent on a C,, solid phase extraction

(SPE) column and not change a complex mixture such as

effluents, and arrive at some false conclusions about the

toxicants in the earlier phases.

Not all approaches discussed in the following

sections will be applicable to every effluent, and additional approaches might need to be developed during the

TIE. The various approaches need not be performed in

any particular sequence, and the list of possible approaches will get larger as experience is gained. To

effectively evaluate effluent samples from one particular

discharger to obtain a correlation, substantial calendar

time could be required and any steps for correlation

should be initiated at the beginning stages of Phase III.

Judgement must be made as to how many of the approaches described in Phase III confirmation should be

used and how many samples for each should he CXKR

pleted. How completely Phase III confirmation is done will

determine the authenticity of the outcome. The amount of

confidence in the results of the TIE that is required is

dependent at least in part on the significance of the

decision that will be based on the results. For example, if

a suspect toxicant can be removed by pretreatment or by

a process substitution, a higher degree of uncertainty

may be acceptable than if an expensive treatment plant is

to be built. Such considerations are subjective and cannot

be reduced to a single recommended decision making

process with a specified number of samples.

Time and resources might be conserved if identification (Phase II) and confirmation (Phase Ill) can be

started on the very first effluent sample used in the Phas’e

I characterization. However, this is only possible when the

results from the Phase I characterization are definitive

enough to allow the investigators to proceed to identification and confirmation. In the acute Phase III confirmation

document (EPA, 1989D), although perhaps not explicitly

stated, performing Phase I characterizations on several

samples before attempting Phases II and III was implied.

Initiating the Phase Ill confirmation steps earlier in the TIE

is often particularly useful. In addition, many regulatory

agencies have adopted a policy that requires that the

previous TIE approach be modified. For some dischargers, action might be required after the first exceedence in

toxicity, which means that each effluent sample collected

for toxicity testing is of equal regulatory concern when the

toxicity is greater than the permit allows. This regulatory

practice was not in place in 1989 when the earlier TIE

guidance was available (EPA, 1989D) and at that time we

did not expect that the cause of toxicity in one sample

could be sufficiently deduced as we have been able to do.

The importance of confirmation on several samples is not

reduced by the importance of conducting confirmation

steps on single samples; rather, the cause of toxicity for

each sample must be confirmed.

In addition to the importance of each sample with

toxicity greater than the allowable amount specified in a

permit, a sample that is quite different from the previous

samples must be evaluated to determine if the data point

must be included in the Phase III correlation final data

analyses. For each effluent sample, the data points must

be explainable. If one sample is quite different than other

samples it can cause the correlation to be less useful;

however, if it can be shown to have a different toxicant the

data point for that sample can be eliminated from the

correlation. For example, suppose five consecutive

samples during a Phase III evaluation exhibited toxicity

that correlated well with a suspect toxicant. Then a sixth

sample exhibits greater toxicity than previous samples

while the measured concentration of the suspect toxicant

is much lower than measurements on previous samples.

In this sixth sample, the greater toxicity is thought to be

caused by a different toxicant. Now in plotting the data for

the correlation (Section 2), the datum point for the sixth

%3XV$p, \4ciIl, r;ryJ, & dW;a? +fi +& -j33rrh3 50-t tire ex’rst’rng

regression and could render the correlation non-significant. If however, when the sixth sample is then subjected

to intensive study using Phase I characterization and

Phase II identification techniques, and if another toxicant

is identified (or even if Phase I only shows that the toxicity

has very different characteristics), datum for the sixth

sample can legitimately be excluded from the correlation.

This preserves the worth of the data for the previous five

samples. In confirmation, every effort should be made to

determine why a particular sample shows different responses in the various TIE steps from other samples.

This is not to imply that multiple effluent samples

need not be subjected to Phase I manipulations, even if

Phase II and/or Phase III are initiated on the first sample.

Most effluent samples tend to be representative of the

routine effluent discharge. However, determining what is

the characteristic discharge for each effluent is important

to the final success and completeness of the TIE.

When Phase III is completed, all results that were

obtained during the TIE should be explainable. Unless the

results make sense for all samples (aside from an occasional aberrant data point) something has been missed or

is wrong. If so, the confirmation is not complete. Many

techniques used in Phase III require keen observations

and extensive or broad knowledge of both chemistry and

toxicology but above all the ability to synthesize small bits

of evidence in a logical sequence is essential. This TIE

work is most effective when scientists interact daily.

l-2
A note of caution. If data obtained on early samples

during Phase I are to be used for Phase III purposes,

quality control will have to be suitable to provide defensible data (cf., EPA, 1991A; EPA, 1992; EPA, 1993A). In

Phases I and II, the permissibility of using small numbers

of animals and replicates, and omitting measurements

such as pH, DO, and temperature that are required for

routine monitoring tests or single chemical tests was

discussed (EPA, 1989E; EPA, 1991 A; EPA, 1992; EPA,

1993A). These modifications were made to reduce cost

and allow more testing, but at this point shortcuts must be

avoided because definitive data that constitute the basis

for important decisions are generated in Phase III. For

Phase III testing, the effluent test protocols that triggered

the TIE (EPA, 199X; EPA, 1993B) should be followed,

paying careful attention to test conditions, replicates,

quality of test animals, representativeness of the effluent

samples tested, and strict QA/QC analytical procedures

including blanks and recovery measurements. Analytical

work must be selective for the identity of the toxicant and

its concentration measurement. When small differences

in toxicity must be detected, concentration intervals should

be smaller to obtain partial effects (e.g., use dilution

factors of 0.60 or 0.65 versus 0.5). Remember, all of the

data from Phases I and II (for either acute or chronic

toxicity) are considered preliminary relative to Phase III

data. However, if a suspect toxicant is identified and

Phases I and II data may be necessary for confirmation,

stricter QA/QC can be applied for each of the subsequent

Phases I and II techniques so that the data can be used in

Phase III.

For samples exhibiting chronic toxicity, modifications or ‘changes to some of the TIE procedures are

required for confirming the cause of chronic toxicity. Remember that for confirmation (as well as for Phases I and

II), only a single sample of effluent should be used for

each renewal in any chronic test (cf., EPA, 1992; EPA,

1993A). This is important because one cannot correlate a

measured concentration of a toxicant with the toxicity

measured in a test if multiple samples are used for each

-

renewal and the toxicant is not present in some samples

but other toxicants appear. Even more likely, the ratios of

the toxicants, when more than one is present, might

change from sample to sample. In these instances, there

is no valid way to calculate the toxicity of a given toxicant.

Overall, considerations for chronic toxicity tests in Phase

III are not much different than acute toxicity tests in Phase

III. At present, permit requirements specify the 7-d test

and unless data are gathered to show that the 4-d and 7-

d tests yield the same results and that the same toxicants

are involved, the 7-d test should be used for confirmation

(cf., EPA, 1993A). If the 4-d Ceriodaphnia dubia test has

been used instead of the 7-d C. dubia test (see EPA,

1992) during Phases I and II, serious consideration should

be given to returning to the 7-d test for Phase III.

When identification of the toxicant causing

chronic toxicity is desired, and the effluent also exhibits

acute toxicity, it might be possible to use acute toxicity as

a surrogate measure to characterize the toxicity in Phase

I and assist in an identification in Phase II. It must be

demonstrated that the cause of the acute toxicity is the

same toxicant as the toxicant causing the chronic

toxicity. Yet for confirmation, use of chronic toxicity endpoints to confirm the cause of the chronic toxicity is

strongly recommended to avoid misleading the TIE results when using acute toxicity as a surrogate for chronic

toxicity. As discussed in the chronic Phase I manual

(Section 5.8; EPA, 1992), effect levels for chronic tests

should be calculated using the linear interpolation method

rather than the hypothesis test (EPA, 1992). In order to

get more precise estimates of endpoints, test concentration intervals might have to be narrowed (see above).

However, when point estimation techniques for other than

survival endpoints (such as the inhibition concentration

(ICp); EPA, 1993B) are used, a point estimate effect

concentration can be estimated. The effect concentration

estimates will also be more accurate when intermediate

concentrations are used (i.e., use dilution factors of 0.6 or

0.65).

1-3
Section 2

Correlation Approach

2.1 Correlation

The purpose of the correlation approach is to

show whether or not there is a consistent relationship

between the concentration of suspect toxicant and

effluent toxicity. For the correlation approach to be useful,

the toxicity test results with the effluent must demonstrate

a wide range of toxicity with several effluent samples to

provide an adequate range of effect concentrations for

the regression analysis. For sediment samples, spatial

variability might be used to perform correlation analyses

(EPA, 19916).

The effluent effect concentration (i.e., LC50 or

ICp) data and the measured toxicant concentration data

must be transformed to toxic units (TUs) for the regression analysis to evaluate whether or not a linear relationship exists. Effluent TUs are obtained by dividing 100%

by the effect concentration expressed in percent of the

effluent (cf., EPA, 199lA; EPA, 1992). The suspect toxicant concentration is converted to TUs by dividing the

measured toxicant concentration by the LC50 or ICp for

that toxicant (data to make this comparison might have to

be generated; EPA, 1993A). If more than one toxicant is

present, the concentration of each one is divided by the

respective LC50 or ICp value and the TUs can then be

summed (cf., discussion below for non-additive toxicants).

Most of the effluents we have tested have exhibited a wide range of toxicity with several different samples

and therefore the data can be used in the correlation

approach. Typically for the correlations that .we have

conducted, the data used are from toxicity tests without

any manipulations and from chemical measurements on

the effluent samples for the concentrations of the suspect

toxicant. However for effluents where ammonia was the

cause of the toxicity, the effluent toxicity. results have not

varied in toxicity enough, nor have the ammonia concentrations fluctuated enough to use the data in a correlation.

Also, when the effect concentration is greater than 1 OO%,

this information is not useful since the data point cannot

be included in the regression analysis. However, when

samples are marginally toxic or when the suspect toxicant

concentrations do not vary enough from sample to sample

(i.e., ammonia is cause of toxicity), changes in toxicity can

be induced by sample manipulation (cf., EPA, 1993A) and

this toxicity data can be used to develop a different type of

correlation. For example, the toxicity of a given amount of

total ammonia can be changed by over an order of

magnitude by altering the pH of aliquots of the effluent

within an acceptable physiological range (e.g, pH 6 to 9).

For some metals and some species, the toxicity can also

be changed by adjusting the pH and using dilution waters

of varying hardness. This type of data is useful in the

correlation step as providing additional weight of evidence. Therefore, the idea of minimal manipulation(s)

and any risk of creating artifactual toxicity are off set by the

utility of the data.

An example of the regression from an effluent

from a POTW in which the suspect toxicant was diazinon

is given in Figure 2-l. The independent variable (x-axis) is

the TUs of diazinon and the dependent variable (y-axis) is

the effluent TUs. The solid line is the observed regression

line obtained from the data points, and the dashed line is

the expected or theoretical regression line. If there is 1 .O

TU of the toxicant in 100% effluent, then the effluent

should have 1 .O TU (i.e., the LC50 =lOO%). Likewise for

2.0 TUs of suspect toxicant, the effluent TUs should be

2.0, et cetera. Thus, the expected line has a slope of one

and an intercept of zero. In Figure 2-1, the intercept (0.19)

is not significantly different from zero and the slope is very

close to 1 (1.05). The P value is 0.63 which, while not

high, indicates that the majority of the effluent toxicity is

explained by the concentration of the toxicant. As the r2

becomes lower, less confidence can be placed on slope

and intercept. In a small data set such as this, one datum

point that had 5.0 TUs for the effluent toxicity lowered the

r2 value substantially. As discussed in Section 1, if an

intensive effort had been expended on that sixth sample

and another toxicant had been found, this particular

datum point could have been excluded and the P value

would have been higher.

In another POTW effluent, diazinon was also the

suspect toxicant. For these data (Figure 2-2) the slope is

1.38, the intercept is 1.24 and the r2 value is only 0.15,

which all indicate poor fit for diazinon as the only toxicant.

The low r2 value indicates a large amount of scatter,

therefore little can be inferred from the slope and the

intercept. Based on this correlation, we returned to Phase

II analytical procedures and identified two other organophosphates (chlorfenvinphos (CVP) and malathion). Toxicity data indicated that CVP was present at toxic

2-l
1 2 3 4 5 6

TUs of Suspect Toxicant

Figure 2-1. Correlation of toxic units (TUs) for an effluent and one

suspect toxicant in POTW effluent.

0

0

0

0

0

0

1.24

Figure 2-2.

1 2 3 4 5 6

TUs of Suspect Toxicarlt(s)

Correlation of toxic units iTUs) for an effluent and cne

suspect toxicanr in a POTW effluent when two toxicants

are the cause of toxicity.

concentrations while malathion was no?. After testing

each compound both separately and as a mixture, the

toxicity from all three chemicals was determined to be

additive, so a new correlation was begun with analytical

measurements made for all three chemicals. CVP and

diazinon have nearly identical LC50 values for the species (C. dub@ used in this TIE. Malathion is about one-

.

fourth as toxic as CVP or diazinon. Since the measured

concentrations of malathion were lower than its toxicity, it

was not included in the regression analysis. In a new

correlation with data for the TUs summed for CVP and

diazinon versus the effluent TUs, the data show a much

better fit to the expected slope and intercept and a high r2

value (Figure Z-3). Malathion TUs could also have been

included in the regression (although its contribution to

toxicity was minimal) because it was additive with other

toxicants. This type of situation is discussed below.

In addition to slope and intercept, some judgement of the scatter about the regression line must be

made. This can be done statistically, but when the sample

size is large, the scatter can be very large and yet not

negate the relationship. A-suggested approach to avoid

the effect of sample size on the significance of scatter is

to set a lower limit on r2. This value (often expressed as

percent) provides the measure of how much of the observed effluent toxicity is correlated to the measured

toxicant. It is not dependent on choosing the correct effect

concentration of the toxicant. The specific choice of the

minimum value of r2 should be made based upon the

consequences of the decision. It is important to recognize

that experimental error makes an r2 value greater th2r-r

0.80 or 0.85 difficult to obtain. Therefore, where minimal

chance of an incorrect decision is required, an r2 value of

nearly 0.80 may be used. Where an increased risk of an

incorrect decision (i.e., a lesser amount of the toxicity

accounted for) is acceptable, a lower value such as 0.60

may be used.

Since <I .O TU cannot be directly measured in the

effluent, such values are, of necessity, excluded from the

regression. (This comment is exclusive of the use of

concentrates such as the C,, SPE fractions’ where TUs of

~1 .O are possible.} However in some instances, when the

TUs based on chemical analyses are ~1 .O TU and effluent effect values are ~1 .O TU, the data support the validity

of the regression provided a suspect toxicant has been

found in several previous samples. In the correlation for

the effluent toxicity depicted in Figure 2-2, toxicity was

present in a different fraction (Phase II non-polar organic

identification) than where the pesticides were identified. A

specific toxicant was not identified in that fraction and

toxicity was not always measurable in that fraction. However, this additional toxicity may have decreased the r2

value.

Correlation might be more definitive when two or

more toxicants are present. For example, suppose three

toxicants are involved. If each toxicant has the same

LC50 and each is strictly additive with the ratio of their

concentrations remaining the same, the slope will be the

expected but the intercept will be positive if all toxicants

ITUs can be calculated from toxicity tests with the fractions, the

concentrate or the HPLC fractions as described in Phase II (EPA,’

1993A).

2-2
.

slope = 0.82

y-intercept = 0.46

4

H

0” ” ” ” ” ”

0 1s 2 3 4 5 6

TUs of Suspect Toxicant

Figure 2-3. Correl?tion of toxic units (TUs) for an effluent and two

toxicants in a POTW effluent.

are not identified. If the relative amounts (ratios) of each

toxicant vary from sample to sample, the slope, intercept

and r2 will be different from the expected if only one

toxicant is identified. If the toxicity of one of the toxicants

is substantially different, and if the ratios of the three

toxicants vary from sample to sample, then the slope,

intercept, and r2 value will all be different from expected if

all are not identified. Much can be learned from studying

the interrelationship of slope, intercept and the r2 value.

For example, a high r2 value and an intercept near zero

with a slope larger than 1 can be caused by using an

effect concentration for the toxicant that is not appropriate

for the toxicant in the effluent matrix (e.g., suspect toxicant is more toxic in effluent matrix than in single chemical test). This error causes the toxicant TUs to be too few

relative to the effluent TUs (Figure 2-4) (cf., discussion

below on non-additive toxicants). If toxicant concentrations and effluent toxicity show a wide distribution, a

significant correlation will be easier to demonstrate than

for a narrow range.

Great care must be taken to understand whether

or not toxicants are additive or if the TUs for each toxicant

are so different that only one toxicant determines the

effect level. For either situation, the resulting data will

have to be interpreted as though the toxicants are nonadditive. For example, suppose the ratio of TUs is so

disparate that at the effluent effect concentration, the

toxicant with fewer TUs is always present at a fraction of a

TU (e.g., 0.25 of a TU). Whether the two toxicants are

additive or not is irrelevant because the major toxicant will

set the effluent effect concentration. While 0.25 TUs of

the minor toxicant appear to be relatively unimportant in

view of experimental variability, this affects the regression. If in one sample the effect concentration is 25% and

the 4 to 1 ratio of toxicants occurs, there are 4 TUs of the

major toxicant and 1 TU of the minor toxicant. If the

toxicant concentrations are summed, 5 TUs will be plotted

against 4 effluent TUs, and this results in a 25% error.

When secondary toxicants are present in concentrations

that will not contribute to the effect concentration of the

effluent, they should not be included in the correlation

data set. Obviously if an effluent had several toxicants in

dissimilar ratios, the error of including the minor TUs in a

correlation plot could be large and may negate the correlation significance. The investigator should evaluate the

data in regression plots to consider the significance of the

contribution of the secondary toxicant especially if the

toxicants appear to be additive.

Unfortunately the minimum fraction of a TU that is

detectable will depend on the precision of the laboratory

performing the testing. And of course the precision of the

testing is not only dependent on the quality of the work,

but the inherent precision of measuring specific toxicant

TUs. That is, the toxicity measurement for some chemicals is more precise than for some other chemicals. In

general, a chemical such as NaCl whose toxicity is generally not affected by pH, alkalinity, hardness, total organic

carbon (TOC), suspended solids or solubility, can be

measured more precisely than a chemical whose toxicity

is affected by these factors, such as lead or copper.

Therefore, each laboratory must determine which fractional value of a TU at the effect concentration is

unmeasurable, thus indicating which TUs contributed by

the minor toxicant should be deleted from the correlation

data set.

Clearly, if two or more toxicants are strictly nonadditive, then only the major one (the one present in the

most TUs) should be included in the correlation data set.

Since additivity might be easier to measure than the

minimum measurable contribution of a fraction of a TU, it

may be preferable to first determine if additivity occurs. If

substances appear to be partially additive, then very

careful work is required to properly add TUs.

Some very unusual decisions are required in

accepting data into the correlation database when toxicants are strictly non-additive. For example, consider zinc

and ammonia in the same effluent sample; we have fcund

them to be strictly non-additive. Also consider that in

some samples zinc and ammonia occur in TU ratios of 3

to 1 and in other samples the ratio is 1 to 2. In the

regression for the 3 to 1 ratio samples, only zinc TUs

should be plotted. In the regression for the 1 to 2 ratio

samples, only ammonia TUs should be plotted. For this

particular example, 3 TUs for the first sample and 2 TUs

for the second sample would be used if the data is

interpreted correctly (i.e., plotting total TUs) or 4 and 3

TUs would be used respectively, if the data is interpreted
4

Correct Plot

I I I I

1 2 3 4

Calculated TUs

Incorrect Plot

1 I I I

1 2 3 4

Calculated TUs

Figure 2-4. Correct (top) and incorrect (bottom) plots of toxic units

(TUs) for non-additive toxicants.

incorrectly. The slopes for both plots would be 1 but a

negative intercept instead of an intercept of 0 would be

obtained for the incorrect plot. The more similar the TUs

of each toxicant are to each other, the greater the error in

the correlation will be.

2.2 Correlation Problems Caused by Matrix

Effects

Correlation becomes much more difficult when

the toxicants interact with the other effluent constituents

in ways that change their toxicity and we refer to these

changes as matrix effecti. There are numerous matrix

effects and all of them will not be discussed here; instead

2-4

.

a framework is provided to aid in designing tests or test

conditions to validly incorporate matrix effects in such a

manner that useable correlation data can be obtained.

Matrix effects generally fit into one of two categories. One category is when the toxicants change form in

some manner which exhibit a different toxicity. A very

common example is ammonia which changes from NH, to

NH,+ as pH decreases. NH,+ is so much less toxic than

NH, that it is often considered nontoxic? Another example

is HCN whose most toxic form is as un-dissociated HCN,

a form predominating at low pH values. As pH increases

the equilibrium shifts to more H+ and CN-. If metals are

present, metal-cyanide complexes form which are often

less toxic than HCN but metal-cyanide complexes might

vary in toxicity depending on the metal. For example, ironcyanide complexes are much less toxic than some of the

other metal complexes. Metal-cyanide complexes might

also photodecompose in sunlight releasing HCN or H+

and CN-, depending on PH.

A second category of matrix effects involves such

physical changes as sorption or binding in some manner

so as to make the toxicant unavailable to the organism.

For example, non-polar organics sorb onto suspended

solids, and some metals, such as copper, also sorb o%o

suspended solids. The presence of organic matter on

suspended solids might increase the sorbtive capacity.

Predictably, changes in water chemistry often change the

sorption/solution equilibrium and thereby, change the portion of total toxicant that is available to the organism.

To further complicate matters, biological characteristics of the test organisms might change the availability of the same toxicant form. For example a non-polar

organic sorbed on suspended solids such as bacterial

cells, might be unavailable to a fish but readily available to

daphnids because cells might be ingested and digested

by daphnids. The uptake route then is through the digestive tract but the toxicant has entered the body none-theless.

From the above discussion, it is obvious that one

method of correlation will not be applicable for all toxicants. A temptation may be to remove the toxicant from

the effluent and then use the effluent as a diluent to

measure toxicity. However, because effluents are so complex and undefined, there is virtually no way to remove

one or a few constituents and still be certain other characteristics have not been changed. For example, zeolite

removes ammonia but it also removes some metals and

non-polar organics; the C,, resin removes metals as well

as non-polar organics; Ion exchange columns remove

ionized constituents, but non-polar organics also are retained by the columns. Toxicant removal procedures have

utility but require very complicated simultaneous testing

of the effluent and proper blanks (cf., EPA, 1992; EPA,

*See specific discussion in Section 3, Phase II (EPA, 1993A).
.

5 0 1993A) is necessary to properly interpret results (cf.,

Section 9 on hidden toxicants).

In Phase III, quantitative comparisons are being

made between toxicity and concentrations of toxicants

rather than qualitative comparisons as in Phases I and II

(EPA, 199lA; EPA, 1992; EPA, 1993A). In the correlation

approach, such comparisons are the essence of the

technique. Therefore even small changes in form or availability might be unacceptable. This means that manipulations and changes must be minimized when effluent

toxicity and toxicant concentrations are to be compared.

Solvent extraction, so commonly used for organic

analyses, is likely to extract biologically unavailable organics as well as soluble forms. The total measured

concentration may be larger than the true exposure concentration. Use of the C, SPE column also is not free

from problems as the C,, SPE column is a finer filter than

the glass fiber filters commonly used for pre-column

filtration. Therefore solids are likely to be physically retained on the upper part of the column. When the column

is eluted with methanol, the methanol extracts toxicant

from the solids (which might not be biologically available)

as well as elutes the C,, sorbent itself. For Phases I and

II, this might be unimportant, but for the Phase III correlation step where careful quantitative comparison is necessary, the effect might be unacceptable. Such problems

probably reach a maximum when working with samples

such as highly organic sediment pore water (with high

organic characteristics) where much of the chemical might

be biologically unavailable.

The central problem for either type of matrix

effect is the difficulty of analytically measuring the biologically available portion of the specific toxic form. A correlation for a POTW effluent where for nickel was suspected

of causing the toxicity is shown in Figure 2-5. During

Phase I, the acute toxicity was removed with EDTA

additions, and in Phase II the nickel was measured at

toxic concentrations to C. dubia. The toxicity correlated

very well with total nickel concentration (r2 = 0.89 and a

slope of 1 .17) and it appeared that only nickel seems to

be involved. But the intercept of -12.34 is quite different

from the expected zero. Such an intercept would be

expected if there were a relatively fixed amount of nickel

which was not biologically available in all samples. In this

example, because all other confirmation data corroborated nickel as the toxicant, a constant concentration of

nontoxic nickel was thought to provide the explanation for

the unexpected intercept value. However, there is no

obvious reason to think that the quantity, or even the

percentage of total toxicant, is the same across samples

for other toxicants, or for nickel in other matrices.

For the effluent samples that lose their toxicity in

a short time, the nontoxic effluent can be used for the

suspect toxicant tests as a diluent in parallel tests

using a standard dilution water to elucidate matrix effects

on toxicity. Toxicity test results with quite different toxicity

would reflect matrix effects. If toxicity is persistent, devel40

10

y-intercept = -12.34

0 10 20 30 40

TUs of Suspect Toxicant (Nickel)

Figure 2-5. Correlation of toxic units (TUs) for a POTW effluent and

the suspect toxicant, nickel.

oping two separate correlations using pure cheplical additions on two different effluent samples, each with substantially different toxicant concentrations, might be useful.

If the toxicity test results indicate that the biologically

unavailable portion changes with measured concentrations, the slope should be different than one. This approach requires careful work and the investigator must

consider incorporating equilibrium time experiments (cf.,

EPA, 1993A).

Metals can be especially difficult toxicants to

implicate using correlation because the toxicity of metals

is typically very matrix dependent. When the knowledge

of these characteristics is extensive for a chemical, as it is

with ammonia (see Phase II), testing can be tailored to

the chemical and a very powerful correlation obtained.

The large amount of available information on ammonia

does not exist for most metals. In these instances, the

logic pattern should to be reversed where the approach

has to become: if x is the toxicant, what are the matrix

effects?. These can be found by pure chemical testing

combined with Phases I or II manipulations. Once an

adequate understanding of matrix effects is obtained, the

information can be used to answer the question: Is the

effluent toxicant behavior consistent with the matrix effects for the suspect toxicant?

Matrix effects will have varying impacts on toxicant behavior that also depends on the effluent effect

concentration. For effluents which have effect concentrations in the ~10% range, the test solutions will more

closely resemble the diluent water matrix than the effluent. If the effluent has effect concentrations in the 50% to

100% range, the matrix effects of the test solution will

most likely resemble those of the effluent, not of the

dilution water. Since effluent TUs are calculated from

2-5

-
responses occurring in the dilution near the effect concentration, the matrix characteristics of that concentration

are of the most concern for correlation. Thus the importance of the effluent matrix effects diminishes as the

toxicity of the effluent is greater (i.e., matrix at effect level

is more like dilution water).

One can safely say that the difficulty of simulating

the matrix effects with a simulated effluent is quite large

so that the choice is clearly to use the actual effluent

when possible. An important reason for this choice is that

so few matrix effects have been studied extensively, and

beyond pH and hardness little data exists. Even then the

interrelationship between pH, alkalinity and hardness were

often ignored.

The above discussion does not provide all of the

options on how to handle matrix effects. However, it

.

should provide convincing evidence that more than the

correlation step alone is necessary to provide adequate

confirmation!

In summary, the TIE research experience has

revealed two major areas of potential problems in using

the correlation approach. The lack of additivity for toxicants found in effluents requires careful analysis when

calculating TUs for regression purposes. Secondly, when

there are matrix effects, correlation becomes difficult because the effluent matrix might change from sample to

sample and because there are no analyses specific for

the toxic forms. For such effluents, other confirmation

techniques should be used more extensively to better

support the overall confirmatory efforts.

c

2-6
Section 3

Symptom Approach

Different chemicals may produce similar or very

different symptoms in a test species. Probably no symptom of intoxication is unique to only one chemical. Therefore, while similar symptoms observed between two

samples means the toxicant could be the same or

different, different symptoms means the toxicant is

definitely different, or there are multiple toxicants in the

two samples. By observing the symptoms displayed by

the test organisms in the effluent and comparing them to

the symptoms displayed by test organisms exposed to

the suspect toxicants, failure to display the same symptoms means the suspect toxicant is probably not the

true one or the only one.

Behavior of most test species is difficult to put

into words so that a clear image of behavior is obtained.

Behavioral and morphological changes of 30-d old fathead

minnows (Pimephales promelas) were used as diagnostic endpoints in 96 h flow-through single chemical tests.

Organic chemicals of various modes of action were tested

and video recordings were used to monitor the behavioral response (Drummond et al., 1986; Drummond and

Russom, 1990). Substances within a single chemical

classification did not necessarily cause the same type of

response (Drummond and Russom, 1990). Therefore, it

is difficult to predict chemical classification using behavioral monitoring alone.

This type of behavioral monitoring data does not

exist for the cladocerans or the newly hatched fathead

minnows or other species that are most frequently used

in the TIE process. However, noting various symptoms is

useful in the TIE. This is done by simply exposing the test

species to the suspect toxicant and observing how

they react. By the time confirmation is initiated, toxicity

tests with the suspect toxicants will have been conducted

using pure compounds and symptoms may have been

observed. It is important to note the symptoms observed

during all testing because such characteristics can be

very helpful in confirmatory work.

The intensity of exposure concentrations might

change the symptoms observed with the suspect toxicant

in the effluent. Therefore, it is important to compare

symptoms at concentrations that require about the same

period of onset. This can be done by comparing symptoms at exposure concentrations that have similar TUs. In

this way both the unknown (sample) and the known

toxicants (pure compound) can be set at the same toxicity

level.

Observations of the organisms should not be

delayed until the normal length of the test has elapsed.

With some toxicants, the test organisms will show distinctive symptoms soon after the exposure begins, whereas

later, symptoms are often more generalized and less

helpful. For some other toxicants, a sequence of different

symptom types are displayed by the test organism over

the exposure period and the sequence may be more

definitive for a given chemical than the individual symptoms. In few cases will the symptoms be unique enough

to specifically identify the toxicant, but symptoms different

from those caused by the pure suspect toxicant are

convincing evidence that the suspect toxicant is not the

true or only one.

A second caution is needed regarding mixtures of

toxicants. Mixtures of toxicants can produce symptoms in

test animals different from the symptoms of the individual

toxicants comprising the mixture. When more than one

toxicant is involved, the investigator must not only include

all the toxicants, but include them in the same ratio as

measured in the effluent. Often the toxicant of the mixture

at the highest concentration relative to its effect concentration will cause most of the symptoms. As for single

toxicants, the mixture concentration causing the same

endpoint in a similar exposure period should be compared. Spiking effluent with the suspect toxicants and

comparing the results of the spiked effluent sample and

the unspiked effluent sample toxicity tests, both near their

effect concentrations, is a good approach to take (Section 5).

Symptoms caused by the toxicant might be

quite different among different species of organisms;

therefore the use of two or more species provides increased definitiveness of the observations. For both species, the researcher must compare symptoms at

concentrations that are equitoxic. The greater the difference in sensitivity, the more important this becomes, The

chemical concentration is unimportant; the important consideration is that equitoxic concentrations are compared.

3-l
Suppose, for example, species A and B have LC50

values for a suspect toxicant of 1 and 80 mg/l. Then

concentrations of 2 and 160 mg/l may be used to compare symptoms of species A and 9, respectively. If the

onset of symptoms is rapid, then perhaps 1.25 and 100

mg/l (1.25xLC50) should be tried. Since symptoms vary

with the exposure intensity, using various multiples of the

LC50 (i.e., 0.5, 1, 2x) can add additional confirmation

data, if the same set of symptoms are seen in both series.

If more than one toxicant is involved, and the ratio of the

two species’ LC50 values for toxicant A is markedly

different than for toxicant 9, C, D, . . . . then the definitiveness of using symptoms is even greater.

For acute toxicity, time-to-mortality at equitoxic

concentrations can be used as a symptom type of test.

Some chemicals cause mortality quickly and some cause

mortality slowly. If for two effluent samples, toxicity is

expressed quickly for one and for the other very slowly,

the toxicants are probably not the same.

In chronic testing, use of symptoms is also applicable. For example, adult mortality, number of young/

female, death of young at birth, growth retardation, abortion, or time to onset of symptoms, all can also be

monitored and such observations may be useful. The

shape of the dose response curve may also be a determinant in assisting in confirmation. Some chemicals show

an all or none type of response (diazinon) while others

(i.e.. NaCI) display a relatively flat concentration-response

slope for chronic toxicity.

-

3-2
Section 4

Species Sensitivity Approach

The effect concentrations can be compared for

the effluent of concern and the suspect toxicants, using

species of different sensitivities. If the suspect toxicant

is the true one(s), the effect levels of effluent samples

with different toxicity to one species will have the same

ratio as for a second species of different sensitivity. Also

the ratio for each species should be the same as for

known concentrations of the pure toxicant. The same

ratio of effect values for two species implies the same

toxicant in both samples of effluent. Obtaining the same

effluent toxicity ratio among various effluent samples for

each species as is obtained by exposure to comparable

concentrations of known toxicants, implies that the suspect toxicants are the actual ones present. However, if

other effluent characteristics affect toxicity and if they

vary, the ratios could also be affected.

The common notion that goldfish are resistant to

most toxicants and trout are sensitive to most toxicants is

not readily substantiated (AQUIRE, 1992). Many species

are more sensitive to certain groups of toxicants than

trout. Of course, there are generalizations that can be

made. For example, sunfish (Centrarchids), frequently

are much more resistant to metals than goldfish, minnows, and daphnids (AQUIRE, 1992). Daphnids tend to

be more resistant to chlorinated hydrocarbon insecticides

than many fish species and more sensitive to organophosphate insecticides (AQUIRE, 1992). These differences must always be verified for the suspect toxicants;

generalities can only be used as an initial guide to

species selection. Sensitivity differences of 1 O-i 00x may

occur in some chemical groups and not in others. If

several toxicants are involved, interpreting the results

and designing the ancillary experiments is more difficult.

If successful, the power of the result for multiple toxicants

is much greater than for a single toxicant. The difference

in sensitivity between Ceriodaphnia and fathead minnows has, on several occasions, revealed either a change

in the suspect toxicants present in a series of effluent

samples, or the presence of other toxicants in addition to

those suspected.

-

_

Comparison of sensitivity among species has

another very important use. Some species may evidence

toxicity from an effluent constituent that the TIE test

species did not. If this happens, then the above comparison will be confused, but at least there will be a warning

that the suspect toxicant may not be the cause of toxicity.

In order to determine what is happening, the investigator

should step back to Phase II, and possibly step back to

Phase I to characterize the additional toxicant and then

identify the toxicant using the new species. A second

Phase III effort might be necessary for this toxicant and

species. It is important not to assume that tb resident

species have the same sensitivity as the TIE test species.

Especially for freshwater discharges into saltwater this

concern is critical when a saltwater organism triggered

the TIE, because at present the techniques and procedures described in Phases I and II are most likely to be

done using freshwater organisms especially since the

effluent is freshwater. If the concern is for marine organisms and their protection cannot be assumed (cf., Section

8, Phase I; EPA, 1991A), confirmation must be conducted

with marine organisms.

In chronic testing, chemical and physical conditions might differ more among tests on different species

because food must be provided during the test period and

different foods are used for each species. For example,

the final pH of fathead minnow 7-d tests might be lower

than in acute fathead minnow tests and both are likely to

be lower than in Ceriodaphnia chronic tests due to greater

respiration rates for fish than cladocerans and food in fish

tests. If the investigation was to confirm ammonia toxicity,

this pH difference could result in confusing results by

showing the Ceriodaphnia to be more sensitive than the

fathead minnows when the reverse should be true (cf.,

EPA, 1993A; Phase II). The above example illustrates

reasons to maintain careful quality control in Phase III

work.

4-l
Section 5

Spiking Approach

In spiking experiments, the concentration of the

suspect toxicant is increased in the effluent sample

and then toxicity is measured to see whether toxicity is

increased in proportion to the increase in concentration.

While not conclusive, if toxicity increases proportionally

to an increase in concentration, considerable confidence

is gained about the true toxicant( Two principles form

the basis for this added confidence. To get a proportional

increase in toxicity from the addition of the suspect

toxicant when it is in fact not the true toxicant, both the

true and suspect toxicants would to have 1) very similar

toxicity and 2) to be strictly additive. The probability of

both of these coinciding by chance is small.

Removing the suspect toxicants from the effluent

without removing other constituents or in some way

altering the effluent is usually not possible. The inability

to do this makes the task of establishing the true toxicity

of the suspect toxicants in the effluent difficult. For many

toxicants, effluent characteristics, such as TOC, suspended solids, or hardness, affect the toxicity of a given

concentration, Some characteristics, such as hardness,

can be duplicated in a dilution water, but certainly not

TOC or suspended solids because there are many types

of TOC and suspended solids, and generic measurements do not distinguish among the different types. For

example, effluent TOC occurs as both dissolved and

suspended solids. In POTW effluents, the source of the

TOC is likely to be largely from biological sources, both

plant and animal (e.g., bacteria) and bacteria are likely to

make up a large component of suspended solids. If there

have been recent storms, oily materials from stormwater

runoff might be high. Simulating TOCs from such variable

sources is next to impossible because TOC is not solely

the result of man-made organic chemicals. For suspended solids, shape, porosity, surface-to-volume ratio,

charge and organic content (all or any), will impact sorption characteristics. None of these qualities are measured by the standard methods for measuring suspended

solids nor can they be reproduced in a simulated effluent.

In a simple system, such as reconstituted soft

water, it is reasonable to expect that for most chemicals a

doubling of the chemical concentration will double the

toxicity, at least in the effect concentration range. If the

solubility of the toxicant is being approached or there are

effects from water characteristics such as suspended

solids, then the toxicity might not double or conceivably

could more than double. For example, if a chemical with a

large n-0ctanoVwater partition coefficient (log P) is largely

sorbed on solids, doubling the total concentration might

more than double the toxicity because the added chemical might remain in solution. Another important issue is

that equilibrium might not be established during the entire

test period and is probably unlikely to occur before the

test organisms are added. For example, in our TIE research, we found various surfactants sorb to solids and

can be removed by filtration (Ankley et al., 1996). In these

experiments, however, filtration failed to remove surfactants immediately after they were spiked in an effluent but

surfactants were removed after a few days equilibrium

time. Other chemicals are likely to show similar behavior

in regard to equilibrium time.

If several toxicants are involved, then their interaction (additivity, independent action, synergism) must be

measured or otherwise included in the confirmation process (cf., Section 2). Since ratios might be as important

as concentration, the best way to spike when multiple

toxicants are involved is to increase each toxicant by the

same number of TUs (e.g., by doubling each). In this way

the ratios of the toxicities remain constant.

The fact that two or more toxicants fail to show

additivity is useful evidence in confirmation. Interpreting

spiking data might require a very high level of competence in both toxicology and chemistry; otherwise the

data could be very misleading. Using more than one

species of differing sensitivity is effective in adding confidence to the results. When matrix effects are complicated, other types of spiking can be done to reduce the

effects of the effluent matrix characteristics. If a method

exists for removing the toxicants from the effluent, such

as the C,, SPE procedures (EPA, 1993A), the extracts or

methanol fractions can be spiked with pure chemicals in

addition to spiking effluent, using the same principles as

described for effluents. The advantage in this approach is

that matrix characteristics such as suspended solids and

TOC will be absent or much reduced and will not affect

spiking experiments as much. The disadvantage is that

proof that the extracts or fractions contain the true toxicants must be generated. Some approaches for doing

5-1
,

this are given in Section 6. The use of the spiking approach is especially applicable to fractions from the C,,

I SPE column or the high performance liquid chromatography (HPLC) column used for the isolation of non-polar

organics. In these procedures, the constituents are separated from much of the TOC, suspended solids and

hardness, so that spiked additions might be strictly additive where they might not be in the effluent. Suggestions

and precautions about ratios and all other previously

discussed concerns apply here too. In addition, concerns

about the methanol percentages in the toxicity tests, the

amount of SPE or HPLC eluate required for the toxicity

tests and the issue of toxicity enhancement by methanol

must be considered in order to generate the appropriate

toxicity data. Spiking the methanol fractions with suspect

toxicants, however, does not provide the same confidence about the cause of toxicity in the effluent as spiking

the effluent directly. The mass balance approach described in Section 6 could be coupled with spiking the

effluent with a portion of the fractions to make the data

more relevant to whole effluent toxicity.

For chronic testing spiking a portion of the methanol fractions, such as C,, SPE methanol fractions into

dilution water to mimic the effluent, requires some special

considerations as discussed in the chronic Phase I (EPA,

1992) and the new Phase II (EPA, 1993A). For any test

species, the effects of the methanol at the effluent spiking

concentration for the test species must either be essentially non-existent or clearly established so that proper

interpretation is applied. The use of spiking for chronic

toxicants of the methanol fractions is not as easy as the

spiking for acute toxicants due to the limitations in the

quantity of methanol that would be added with each

fraction for the toxicity test. If the chronic toxicity effect

level is around or ~25% effluent and the highest fraction

tested is 4x higher than the chronic effect level, add-back

tests can be conducted similar to the acute add-backs but

the quantity of methanol required for the testing and

analysis must be considered (cf., Section 2; EPA, 1993A).

As discussed in Phase II, once a suspect toxicant has

been tentatively identified, the steps of confirmation should

be started although sample volumes of methanol eluates

might limit the amount of testing (see Phase II, Secticn 2;

EPA, 1993A) with chronically toxic samples. Spiking of

appropriate levels for chronic toxicity for single chemicals

(or mixtures) is limited as sublethal data are not as

plentiful as acute data. The acute toxicity of some chemicals might be altered by methanol (i.e., surfactants). The

possibility that this is occurring must be checked and a

correction applied if warranted. Spiking fractions also has

applicability for hidden toxicants; refer to Section 9 for

further details.

Spiking can also be done effectively when the

suspect toxicant of concern can be removed. However,

since other toxicants might also be removed, the data

must be carefully interpreted. Ammonia is a good example (cf., Phase II; EPA, 1993A) to use with this technique where one toxicant can be removed. Ammonia can

be removed from the effluent by passing samples over

the zeolite resin, after which the concentration can be

restored in the post-zeolite effluent by the addition of

ammonia. If toxicity is also restored, then it is likely that

there is sufficient ammonia to cause the toxicity observed.

However, it cannot be concluded from these data alone,

that ammonia is the cause of toxicity because the zeolite

can also remove substances other than ammonia. Another substance which is non-additive with ammonia yet

present at a lesser or the same number of TUs could

cause the initial effluent toxicity but not be discernable by

this removal technique. This is an example of a hidden

toxicant (see Section 9). For acute toxicity, zinc could

behave exactly this way because it is non-additive with

ammonia yet zinc is also removed by zeolite. Using other

ammonia removal methods, such as high pH stripping,

followed by spiking to the initial ammonia concentration

will enhance confidence that a hidden toxicant is not

present. Other examples involving the C,, SPE column

and various ion exchange resins would be approached

and interpreted similarly.

5-2
Section 6

Mass Balance Approach

This approach is applicable only to those situations in which the toxicant can be removed from the

effluent and recovered in subsequent manipulation steps.

The objective is to account for all toxicity to assure that

small amounts of toxicity are not being lost. This concern

is partly covered by the correlation approach (Section 2);

however, a totally different toxicant present at a small

concentration could appear as experimental variability in

the correlation and go unnoticed.

The mass balance concept is best described by

illustration for acutely toxic effluents and the C,, SPE

fractions. As described in Phase II (Section 2.2.7; EPA,

1993A) for acutely toxic effluents, the effluent has been

passed over a C,, SPE column which is then eluted with

the methanol/water fractions. After the toxicity tests on

the individual fractions are completed, add-back tests

can be initiated to determine whether all of the toxicity in

the original sample was accounted for in the SPE fractions. For this step, there are three separate tests (with

dilutions and replicates to calculate effect endpoints) that

must be conducted which consist of the all-fraction test,

the toxic-fraction test, and the nontoxic-fraction test. Assuming a complete recovery of all non-polar organics

from the SPE column, this should yield a solution of nonpolar organic compounds equal to the original sample

concentrations, In the mass balance approach, these

add-back tests are conducted using an aliquot of the

effluent that has passed through the C,, SPE column

(post-SPE column nontoxic effluent) or an aliquot of

dilution water. Each toxic fraction is added back to the

post-SPE column effluent, so that each is present at

original effluent concentrations (i.e., lx effluent concentration). For example for acutely toxic effluents, the toxicfraction test solution is prepared using methanol

concentrations as described in Phase II (i.e., Section

2.2.7; EPA, 1993A) and for each fraction where toxicity

was observed in the fraction toxicity test, 30 ul of each is

added to the same IO ml of nontoxic post-C,, SPE

column effluent (or dilution water). A portion of each of

the remaining fractions where toxicity was not demonstrated are now added to a second post-SPE column

aliquot at effluent concentrations for the nontoxic-fraction

test. Finally portions of all the fractions (e.g., n= 8 for

acutely

column

toxic’ effluents) are added to a third post-SPE

aliquot at effluent concentrations for the all-fraction test. If all the toxicity is exhibited in the toxic-fraction

test, then the all-fraction test results and the toxic-fraction

test results should be the same as in the unaltered

effluent. Results from the nontoxic-fraction test should

indicate that no toxicity is present. This mass balance (or

add-back) approach allows the researcher to ascertain

whether or not the toxicity in the toxic-fraction test equals

the effluent toxicity. Small amounts of toxicity can be

undetectable in the toxic-fractions when tested separately

or the toxicant might not have been eluted from the C,,

SPE columns. Unless mass balance experipents are

conducted, such loss of toxicity might not be detected. In

the effluent example discussed in Section 2, the toxicity

was contained usually in the 75%, 80%, and 85% fractions and occasionally in the 70% fraction! The r2-value,

slope, and intercept were all close to the expected values

if two toxicants (diazinon and CVP) were causing the

effluent toxicity (Figure 2-3). However, in Table 6-l the

results of mass balance tests indicate that toxicity from

the all-fraction test was greater than the toxicity of the

toxic-fraction test. While this difference is small, it did

seem to be real and was attributed to a small amount of

another toxicant in the 70% fraction. In 11 of 12 samples,

the results from the all-fraction tests indicate there was

greater toxicity than was found in the toxic-fraction tests.

On the few occasions when the 70% fraction was toxic, it

did not contain any of the three suspect toxicants. Without

the mass balance data, consistent presence of the additional toxicant would not have been discovered.

At the stage where the toxic-fractions have been

identified, the test of the fractions in a mass-balance test

is highly desirable. For chronic toxicity testing, the amount

of eluate available might be limited following the fraction

toxicity tests. Using eluate for the add-back tests might be

a trade-off between tracking toxicity and having sufficient

eluate to concentrate for further analysis. This limits the

add-back tests broad applicability for chronic toxicity TI Es

unless the effluent is toxic enough that at 4x the chronic

effect level, the methanol concentrations do not exceed

3During development of the non-polar organic procedures, various

elution profiles were used that included the 70% methanol/Water

fraction.

6-l
Table 6-1. Comparison of Eff Lent Toxicity and Toxicity Measured in

Effluent Fraction Add-back Tests

Sample Effluent

12/03/87 1.18

01/l 2/88 2.00

01/13/88 1.93

02/03/88-P cl .oo

02/03/88-II 2.00

03/03/88-la 1.15

03/03/88-II 1.33

03/23/88-l 3.70

03/23/88-H 2.86

04128188 2.27

05/ 17188 2.27

05/l 7188 2.27

Toxic Units (TUs)

All-fractions

1.64

2.94

2.86

1.15

1.75

1.06

1.52

3.03

2.86

1.72

2.04

1.67

Toxic-fractions

1.43

3.13

2.53

<I .oo

1.64

cl .oo

1.13

2.86

2.44

1.64

2.00

1.59

Mean 2.13 2.18 2.00

aValues excluded from mean calculations due to less than values.

the organisms tolerance. For chronically toxic samples,

the all-fraction add-back test with C. dubia is not possible

due to high methanol concentrations in test cups unless

chronic toxicity is below 25% and add-backs are done

using 25% effluent as the high test concentration (cf.,

Phase II; EPA, 1993A). The data from the individual

methanol/water tests may be summed; however this approach must be considered more tentative than add-back

tests (see below).

A deficiency in the above approach to mass

balance is that there can be some toxicity in the post-SPE

column effluent which has not been removed by the C,,

SPE but which is not present in concentrations high

enough to detect. The above mass balance approach

alone will not identify this. However, if the add-back tests

described above are repeated using a standard dilution

water, residual toxicity in the post-SPE column effluent

should cause the toxic-fraction test and all-fraction test to

show more toxicity when added to the post-SPE column

effluent than when added to dilution water. A confounding

effect of this approach is that if the toxicity is changed by

matrix effects (suspended solids or TOC), then the toxicity will be different in the clean water test. Matrix effects

can be discerned, in part, by a third spiking experiment

where a portion of all of the fractions and a portion of each

toxic-fraction test are spiked into whole filtered effluent

(which has not passed through the C,, SPE column). If

the addback tests in dilution water indicates greater toxicity than the addback tests with the post-SPE column

effluent, and the same type of addback test experiment

with filtered effluent (i.e., 1 lrn filter) indicate that the

fractions are exactly additive, then matrix effects are

indicated.

Some post-SPE column effluent samples develop

fungal or bacterial growth or perhaps a precipitate forms

after the effluent passes through the column. For the

fungal type of growth, this is thought to occur when some

methanol bleeds into the effluent as it passes through the

column and more rinsing will not eliminate this problem.

Some effluents consistently develop this type of growth in

the post-column effluent while others exhibit this pattern

in only an occasional sample. To alleviate this problem,

conditioning the column with acetonitrile has helped (cf.,

the acute Phase I (EPA, 1991A) and chronic Phase I

(EPA, 1992) for details). When methanol fractions are

spiked into the effluent this problem might or might not be

enhanced; we have found this to be an effluent-specific

occurrence.

Caution is warranted in situations where toxicity

is contained in more than one SPE fraction. The researcher should not necessarily expect the toxicity expressed by each individual fraction that is tested separately

to add up to the total effluent toxicity. First, toxicants may

not be additive and second, some toxicity which cannot

be detected in individual fractions may add to the whole

toxicity. For example, any one C,, SPE fraction may not

show toxicity but may contain some of the toxicant that is

in the adjacent toxic-fraction. In this case, the toxicity of

the toxic-fraction test would be less than expected. If this

happens in more than one pair of fractions, the sum of the

toxicity from the toxic-fraction test will be less than the

effluent toxicity or all-fraction test. These concerns are

especially important when several toxicants are involved

and one or more occur in more than one fraction.

For effluents where the C SPE column is not

used, but where the toxicants can%e removed from the

sample, the same objectives should be achievable, but

the methods will be different. For example, if an effluent

appears to contain a volatile toxicant, the mass balance

could be done on the trap and on the purged sample.

Since we have not yet done mass balance on samples

such as these we have no experience from which to offer

additional guidance or advice.

Some of the mass balance process begins in

Phase II, and there is a subtle difference in the purpose of

mass balances in Phases II and III. In Phase II, usually

only a few samples are used and mass balances are

necessary to determine the need for more identification in

those few samples. The mass balance is useful in early

stages of Phase II as well before toxicants are identified

at all, because it allows the investigator to decide if the

toxicants present at 2x or 4x whole effluent concentrations are also expressing toxicity at lower concentrations.

In Phase III as many samples are tested, the

mass balance approach can provide information over

time with many samples whether or not the suspect

toxicants consistently account for all or the majority of the

toxicity. As illustrated above, the power of the mass

6-2
balance approach to detect small degrees of toxicity is

better than for the correlation approach.

from the sample does not remove biologically non-available portions. An example of this situation may be the

When a portion of the toxicant is not biologically

alternative solvent extraction procedures which may reavailable and therefore does not contribute to toxicity,

move a bound toxicant sorbed on suspended solids

with the solvent and is now toxic, yet it was not toxic in the

care must be taken to assure that removal of the toxicant unaltered sample.

6-3 -
Section 7

Deletion Approach

In some situations, particularly for industrial discharges, keeping the suspect toxicants out of the waste

stream influent or effluent for short periods of time and

also conducting toxicity tests on the wastewater simultaneously may be practical. When this approach can be

used, it offers the most convincing evidence obtainable

that the suspect toxicants are the true ones. Care must

be taken however, that other substances are not deleted

or that some characteristic such as pH does not change

also. If a researcher can be certain that all changes are

known, then this approach is definitive. Changes in the

toxicants with time are as much of a concern here as in

any other approach. These can be handled by the approaches outlined in earlier sections and the deletion

approach need not be done repeatedly; however, if it

were practical to do so, it would certainly be effective. If

some samples do not contain one or more suspect toxicants, these effluent samples can be used to the advantage in confirmation in much the same way as intentional

deletions described in this section can be used to confirm

toxicity.

*

7-1
-

_

Section 8

Additional Approaches

-

This section mentions only a few of many steps

that can be used to further confirm the cause of toxicity.

The steps mentioned are mostly those that we have used

and found helpful and practical.

The pH is one of the most important effluent

characteristics that changes toxicity. The pH of POTW

effluents, sediment pore or elutriate waters, and ambient

waters will almost always rise when they are exposed to

air, especially in the small test volumes used in TIE work.

Commonly, pH in an effluent sample at 25°C will .rise

from 7.1-7.3 to 8.3-8.5 during a 24 h period. That pH

change is enough to increase ammonia toxicity (based

on total ammonia) about three fold. Such pH changes

can destroy work for some purposes, but by regulating

these pH changes, the pH fluctuations can be used to

great advantage for other purposes.

Phase II (EPA, 1993A) describes the use of pH

change to identify ammonia toxicity. The toxicity of some

metals, hydrogen cyanide and hydrogen sulfide among

others, is altered by pH change. Other characteristics,

such as hardness, can also be varied to see if the

changes in toxicity follow a predictable pattern. The

toxicity of some metals could be approached in this way.

Not all equilibria are as rapid as the ammonia equilibrium,

so the amount of time for equilibria to occur should be

controlled and standardized (cf., Phase II; EPA, 1993A).

Various time . periods may have to elapse before the

expected changes occur and this may differ with each

effluent. With the improved methods of pH control described in the Phase I documents (EPA, 1991; EPA,

1992), much more use can be made of pH manipulation.

Often chemicals in effluent samples may not be

biologically available, and if they are not, then they are

not likely to cause toxicity. They may be made biologically available through some manipulation in Phase I and

subsequently identified in Phase II. Through confirmation, the toxicity due to such a toxicant will become

apparent when the correlation indicates a poor fit (cf.,

Section 2). For many toxicants, biological availability can

be demonstrated by measuring body uptake. If the constituent of concern enters the body from the effluent, it is

certainly biologically available. Exposure to pure compounds may be necessary to establish which particular

organ should be evaluated for the toxicant. In acute metal

exposures using fish, most metals concentrate first in the

gills while non-polar organics concentrate in fatty tissues

such as the liver. When a chemical is metabolized by the

organism, a residue measurement for that compound is

not a valid measure of the lethal body burden because it

is unknown whether the metabolite is more of less toxic

than the parent compound. If the suspect toxicant has a

known mode of action, such as the acetylcholinesterase

inhibition produced by organophosphate pesticides, this

exposure effect can be measured to assess if toxic effects

conform with the predicted effect. The use of enzyme

blockers such as piperonyl butoxide (PBO) is also an aid

in confirming toxicity caused by specific classes of toxicants (cf., Phase II: EPA, 1993A).

As additional steps are needed for confirming the

cause of toxicity, combinations of various Phase I and

Phase II procedures should always be used whenever

practical. When several results are combined and all

results are indicating the same type of toxicant, the data

are more conclusive than when only one procedure yields

predicted results.

Total dissolved solids (TDS) are a common problem in certain areas of the country and for certain industries. TDS will not cause toxicity from osmotic stress (this

can easily be shown because their toxicity is not related to

osmotic pressure) but rather TDS acts as a set of specific

toxicants. For toxicity caused by TDS, the ratios and

concentrations of the major cations and anions can be

measured analytically. A similar mix of these major ions

can be added to a dilution water to see if the expected

toxicity is present. By testing various mixtures, the researcher can ascertain which of the TDS components

contribute most to the toxicity.

8-7
Section 9

Hidden Toxicants

In the previous section, references were made to

the problem of hidden toxicants. Essentially there are two

situations which may produce the problem of hidden

toxicants. The first situation occurs when disparate ratios

of TUs of two toxicants are present in the effluent sample.

Since the effect concentration is measured by diluting the

effluent, when disparate ratios occur, the TUs of the

toxicant present in fewer TUs in 100% effluent are so low

at the effect diluent, that its contribution if any, is not

measurable. This problem exists whether the toxicants

are additive or non-additive. This situation generally will

not be encountered in effluents that have very slight

toxicity (i.e., effect concentration 75% to 100%) because

little or no dilution is required to achieve the effect concentration. For those toxicants present in disparate ratios

in effluents with marginal toxicity, the chemical present at

the low levels may be nontoxic even in 100% effluent.

The second situation where hidden toxicant

occurs is when the toxicants are non-additive or partially

additive in the effluent sample. These toxicants may

occur at approximately equal TUs or at disparate ratios of

TUs, as long as those present at lesser TUs are present

at 1 TU in the 100% effluent (cf., discussion of performing

correlation on these types of toxicants, contained in

Section 2).

If confirmation is being conducted for both acute

and chronic toxicity or if acute toxicity is being used as a

surrogate for chronic toxicity, the acute to chronic ratio

must also be considered. For example, consider an

effluent with toxicants A and B for which the acute-tochronic ratios are 3 and 12, respectively and the TUs for

acute toxicity are 2 and 1 in an effluent sample for A and

B, respectively. By definition, 1 acute TU (TU,) for toxicant A equals 3 chronic TUs (TUJ and for B, 1 TU, = 12

TU,. In this example, the acute toxicity of the effluent will

be determined by A and the chronic toxicity will be

determined by B. If in another situation, the acute-tochronic ratios for two compounds were similar, then one

of the toxicants would determine the effect concentration

for both acute and chronic toxicity. These examples

illustrate the importance of acute-to-chronic ratios for

non-additive toxicants. Acute-to-chronic ratios have special importance for additive toxicants when acute toxicity

is being used as a surrogate measure for chronic toxicity.

If acute toxicity is being used as a surrogate it must be

demonstrated that the cause of the acute toxicity is the

same as the chronic toxicity. When acute toxicity is used

as a surrogate for chronic toxicity in Phases I and II,

interpretation of the results can easily be biased and

these considerations are important.

When a toxicant can be removed from the effluent and recovered, the identification of the presence of a

hidden toxicant is more readily known. For example, the

use of the C,, SPE column may remove hidden toxicants.

The toxicant is recovered in the eluate a&measured

both analytically and toxicologically. This type of hidden

toxicant may be observed if ammonia is present at concentrations that could cause toxicity. For example, in an

effluent sample ammonia is present at 3 TUs. Ammonia

will not be removed by the C,, SPE column and yet an

additional 1.5 TU of a non-polar organic toxicant is evident when the C,, SPE eluate test is conducted. If the

discharger applied remedial treatment they would be able

to remove the ammonia toxicity yet the effluent would still

be toxic. The same concept of hidden toxicants can be

found when toxicants are removed by sublation which is

followed by recovery and concentration of toxicity (cf.,

Phase I; EPA, 1991 A; EPA, 1992). For example, sublation

can separate some surfactants, resin or fatty acids, and

polymers from such constituents as metals and ammonia.

Hydrogen sulfide can be removed by a purge and trap

method, thereby separating it from other effluent constituents.

Specific blockers of toxicity such as EDTA for

metals and PBO for organophosphates are also useful in

establishing the cause of toxicity. The more specific the

blocker, the more definitive are the results, However,

present knowledge does not allow us to be certain that

compounds such as EDTA do not also affect the toxicity

of other chemicals. Use of two specific blockers such as

EDTA and sodium thiosulfate for copper, allows more

definitive conclusions (cf., Phase I; EPA, 1992).

Manipulating characteristics such as pH is useful

but can easily mislead thinking. For example, if the effluent has ammonia toxicity, the toxicity due to ammonia

should disappear if the pH is lowered appropriately. These

results do not allow a conclusion that there are no hidden

9-l
toxicants. If, however, the pH is lowered so as to eliminate

ammonia toxicity but the effluent toxicity exists or even

increases, then the likelihood of a hidden toxicant is high.

Unfortunately a complication to this rationale is that the

toxicity expressed at the lower pH may be totally artifactual due to mechanisms of pH adjustments.

The best approach to find hidden toxicants is to

first use, those methods that alter the effluent the least,

can remove and recover removed hidden toxicants, and

are most specific for a few toxicants. This advice is most

applicable where the effort is to try to find out if some

specified type of toxicant is a hidden one, e.g., is there a

non-polar organic as a hidden toxicant.

If, however, the search is for any type of hidden

toxicant then every conceivable technique should be used

that would help to distinguish a hidden toxicant from the

suspect toxicant( Hidden toxicants are very hard to find

when ammonia is the primary toxicant. Various tests used

to identify ammonia as the toxicant, i.e., use of the zeolite

resin, graduated pH tests and air-stripping (EPA, 1993A),

all have a reasonable probability of changing the toxicity

of many other potential toxicants. For instance, it is known

that zeolite removes some non-polar organics and metals. Air-stripping (at pH 11) could also remove or destroy

many other chemicals as it often must be done for a

extended period of time to achieve good ammonia removal. The graduated pH test results might also implicate

a metal as a toxicant (EPA, 1993A). If these tests were

conducted in Phase II (E?A, 1993A) and the results

consistently indicated ammonia toxicity, these .data indicate that there are no hidden toxicants. The required

characteristics for a hidden toxicant to behave exactly as

ammonia are very specific and obtaining results like those

described above for a toxicant other than ammonia is

unlikely.

,

If the hidden toxicant is additive with the suspect

toxicant but occurs in a disparate ratio, the confirmation

effort must first emphasize confirming the cause of toxicity (or remove the toxicity) of the primary toxicant. Then

toxicity from the hidden toxicant should be measurable.

The probability a hidden toxicant that has additive toxicity

will not express its toxicity using several Phase I or Phase

II techniques is less than the probability that a nonadditive toxicant will express its toxicity using several of

the same techniques.

If the remedial action for a primary toxicant is

specific and easy, such as a product substitution, the

search for hidden toxicants perhaps should be done after

the remedial action has reduced or eliminated the primary

toxicant from the effluent. The remedial action (especially

if it is treatment) may also eliminate the hidden toxicant(

What must be avoided if at all possible, is to carry out

expensive remedial action only to find that the effluent is

still toxic.

The problem of hidden toxicants is a major reason a researcher should not accept the presence of toxic

concentrations of suspect toxicant as sufficient confirmation (cf., Section 1). The presence of biologically unavailable forms (cf., Section 8) is a compelling reason not to do

so.

A thorough confirmation is resources well spent

in most instances. Non-additivity and disparate ratios

complicated by non-availability occur too frequently to bypass confirmation. Seasonal changes or changes without

a pattern, in effluent toxicants are further reasons to

perform the confirmation over a period of time to assure

that the entire suite of toxicants has been found.

9-2
Section IO

Conclusions

Often the most laborious and difficult part of the

TIE is developing data to adequately establish the cause

of toxicity. In our experience, frequently the suspect

cause of toxicity is found without difficulty but developing

a convincing case to prove that the suspect cause is the

true toxicant is the challenge.

Especially for POTW plants, this confirmation

phase must be performed over a considerable period of

time to be certain that the cause of toxicity is not changing. TIES on POTWs and some industrial categories are

not likely to be a one time event but will have to be

repeated as long as the inputs to the plant change. Our

current wastewater treatment plants were not designed to

remove specific chemicals, so there is no reason to

expect that they will remove everything which they receive. Especially where the control over the influent is not

complete, as is the case with POTW plants, a solid case

must be developed to assure that the cause of toxicity is

not changing.

10-l ,
Section 11

When the Treatability Approach Has Been Used

As discussed in Phase I, two main approaches

may be used to remove a toxicity problem--toxicant identification and source control or treatability. Phases I and II

involve the first approach while treatability procedures

accompanied by toxicity testing are used in the second

(EPA, 19898; EPA 1989C).

In the second approach, treatment methods are

varied to determine which will remove toxicity without

identifying the specific toxicants. The treatability approach

requires as much confirmation as the toxicant identification approach. Since the treatability approach should

remove toxicity, the confirmation procedures are somewhat different.

Repeat samples should be tested to ensure that

toxicity has been successfully removed. This should be

done over a sufficient length of time to assure that the

range of conditions are included during the confirmation

phase. Such events as seasonal changes, production

changes, storms, and intermittent operations all should

be included during the confirmation phase. Toxicity should

be consistently removed or appropriately reduced, as

required. Either acute or chronic toxicity removal can be

confirmed this way.

One must be absolutely sure that the toxicity to

resident species has been successfully removed. As has

been pointed out in Phases I and II, the effluent constituents producing toxicity to one species may not be the

same for other species. Toxicity by a given treatment

method may remove all toxicity for one species but not for

another. The species of concern must be tested in the

effluent from the treatment method selected. If chronic

toxicity is the concern, this testing may be more difficult

because chronic testing methods may not be available for

resident species. In selected cases, symptoms may be

substituted for the usual endpoints of chronic tests but

their use would be case specific.

11-l
Section 12

References

Ankley, G.T., G.S. Peterson, M .T. Lukasewycz, and D.A.

Jensen. 1990. Characterization of Surfactants in Toxicity Identification Evaluations. Chemosphere 21(1-2):3-

12.

AQUIRE. 1992. Aquatic Toxicity information Retrieval

Database and Technical Support Document. Environmental Research Laboratory, Duluth, MN.

Drummond, R.A., C.L. Russom, D.L. Geiger, and D.L.

DeFoe. 1986. Behavioral and Morphological Changes

in Fathead Minnow (Pimephales promelas) as Diagnostic Endpoints for Screening Chemicals According to

Mode of Action. Aquatic Toxicology and Environmental

Fate: Nineth Volume, ASTM STP 921. T.M. Poston and

R. Purdy, Eds., American Society for Testing and Materials, Philadelphia, PA. pp. 415-435.

Drummond, R.A. and C.L. Russom. 1990. Behavioral

Toxicity Syndromes: A Promising Tool for Assessing

Toxicity Mechanisms in Juvenile Fathead Minnows.

Environ. Toxicol. Chem. 9: 37-46.

EPA. 1988. Methods for Aquatic Toxicity Identification

Evaluations: Phase I Toxicity Characterization Procedures. EPA-600/3-88-034. Environmental Research

Laboratory, Duluth, MN.

EPA. 1989A. Methods for Aquatic Toxicity Identification

Evaluations: Phase II Toxicity Identification Procedures.

EPA-600/3-88-035. Environmental Research Laboratory, Duluth, MN.

EPA. 1989B. Toxicity Reduction Evaluation Protocol for

Municipal Wastewater Treatment Plants. EPA/600/2-

88/062. Water Engineering Research Laboratory, Cincinnati, OH.

EPA. 1989C. Generalized Methodology for Conducting

Industrial Toxicity Reduction Evaluations (TREs). EPA/

600/2-881070. Water Engineering Research Laboratory, Cincinnati, OH.

EPA. 1989D. Methods for Aquatic Toxicity Identification

Evaluations: Phase III Toxicity Confirmation Procedures. EPA-600/3-88-036. Environmental Research

Laboratory, Duluth, MN.

EPA. 1989E. Short-Term Methods for Estimating the

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12-l *U.S.COV&RNMENT PRINTING OFFICE:1993 -7SO-002/80298