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49
Sibley et al. (1996) reported similar results from a 56-d life-cycle test conducted
with the freshwater midge Chironomus tentans exposed to zinc-spiked sediments (Table 3-2).
The test was initiated with newly hatched larvae and lasted through one complete generation
during which survival, growth, emergence and reproduction were monitored. In sediments
where the molar difference between SEM and AVS was <0, at dry wt zinc concentrations as
high as 270 mg/kg, concentrations of zinc in the sediment interstitial water were low and no
adverse effects were observed for any of the biological endpoints measured. Conversely, when
SEM-AVS exceeded 0, AVS and interstitial water concentrations of zinc increased with
increasing treatments (being highest in surficial sediments) and reductions in survival, growth,
emergence and reproduction were observed. Over the course of the study, the absolute
concentration of zinc in the interstitial water in these treatments decreased due to increase in
sediment AVS and loss of zinc due to twice daily renewals of overlying water.
3.3.2 Colonization tests
Hansen et al. (19965) conducted a 118-d benthic colonization experiment in which
sediments were spiked to achieve nominal cadmium/AVS molar ratios of 0.0 (control), 0.1,
0.8 and 3.0 and held in the laboratory in a constant flow of unfiltered seawater (Table 3-2).
Oxidation of AVS in the surficial 2.4 cm of the control treatment within two to four weeks
resulted in sulfide profiles similar to those occurring in sediments in nearby Narragansett Bay,
RI (Boothman and Helmstetter, 1992). In the nominal 0.1 cadmium/AVS treatment, measured
SEMCd was always less than AVS, interstitial cadmium concentrations (<3-10 pg/L) were less
than those likely to cause biological effects, and no significant biological effects were detected.
In the nominal 0.8 cadmium/AVS treatment, measured SEMo, commonly exceeded AVS in the
surficiai 2.4 cm of sediment and interstitial cadmium concentrations (24-157 ^g/L) were
sufficient to be of lexicological significance to highly sensitive species. In this treatment.
shifts in the presence or absence over ail taxa, and fewer macrobenthic polychaetes
(Mediomastus ambiseta, Streblospio benedicti and Podarke obscura) and unidentified
meiofaunal nematodes, were observed. In the nominal 3.0 cadmium/AVS treatment,
concentrations of SEMCd were always greater than AVS throughout the sediment column.
Interstitial cadmium ranged from 28,000 to 174,000 ^g/L. In addition to the effects observed
in the nominal 0.8 cadmium/AVS treatment, sediments in the 3.0 cadmium/AVS treatment
were colonized by fewer macrobenthic species, pplychaete species and harpacticoids; had
lower densities of diatoms; lacked bivalve molluscs; and exhibited other impacts. Over ail
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. 50
treatments, the observed biological responses were consistent with predicted possible adverse
effects resulting from elevated SEM^/AVS ratios in surficial sediments and interstitial water
cadmium concentrations.
Boothman et al. (1996) conducted a field colonization experiment in which
sediments from Narragansett Bay, RI were spiked with an equi-molar mixture of cadmium,
copper, nickel and zinc at nominal SEM:AVS ratios of 0.1, 0.8 and 3.0, placed in boxes, and
replaced in Narragansett Bay (Table 3-2). AVS concentrations decreased with time in surface
(0-3 cm) sediments in all treatments where SEM< AVS, but did not change in subsurface (6-10
cm) sediments or in the entire sediment column in the SEM > AVS treatment. SEM decreased
with time only where SEM exceeded AVS. The concentration of metals in interstitial water
was below detection limits when SEM was less than AVS. When SEM exceeded AVS,
significant concentrations of metals were present in interstitial water in order of their sulfide
solubility product constants. Interstitial water concentrations in these sediments decreased with
time exceeding the WQC in interstitial water for 60 days for all metals, 85 days for cadmium
and zinc, and for the entire experiment (120 days) for zinc. Benthic faunal assemblages in the
spiked sediment treatments were not different from the control treatment. Lack of biological
response was consistent with the vertical profiles of SEM and AVS. AVS was greater than
SEM, in all surface sediments, including the top 2 cm of the 3.0 SEM:AVS treatment, due to
the oxidation of AVS and loss of SEM. The authors speculate that interstitial metal was likely
absent in the surficial sediments in spite of data demonstrating the presence of significant
measured concentrations of interstitial metal. This is because the interstitial water in the
nominal 3.0 SEM/AVS treatment was sampled from sediment depths where SEM was in
excess. It is in surficial sediments where settlement by saltwater benthic organisms first occurs.
Also, there was a storm event which allowed a thin layer of clean sediment to be deposited on
top of the spiked sediment (Boothman. USEPA, persona! communication). These data
demonstrate the importance of sampling of sediments and interstitial water in sediment
horizons \vhere benthic organisms arc active.
Hare et al. (1994) conducted an approximately 1-yr field colonization experiment in
which uncontaminated freshwater sediments were spiked with cadmium, and replaced in the
oligotrophic lake from which they originally had been collected (Table 3-2). Cadmium
concentrations in interstitial waters were very low at cadmium:AVS molar ratios < 1.0, but
increased markedly at ratios > 1.0. They reported reductions in the abundance of only the
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51
chironomid Chironomus salinariits in the nominal 10.0 SEM/AVS treatment. Cadmium was
accumulated by organisms from sediments with surficial SEM concentrations were greater than
those of AVS. These sediments also contained elevated concentrations of cadmium in
interstitial water.
Liber et al. (1996) performed a field colonization experiment using sediments
having 4.46 pmole sulfide/gram from a freshwater mesotrophic pond (Table 3-2). Sediments
were spiked with 0.8, 1.5, 3.0, 6.0 and 12.0 ^mole zinc/gram, replaced in the field and
chemically and biologically sampled over 16 mo. There was a pronounced increase in AVS
concentrations with increasing zinc concentration, AVS was lowest in the surficial CL-2 cm of
sediment with minor seasonal variations. With the exception of the highest spiking
concentration (ca., 700 mg/kg, dry wt), AVS concentrations remained larger than those of
SEM. Interstitial water zinc concentrations were rarely detected in any treatment, and were
never at concentrations that might pose a hazard to benthic raacroinvertebrates. The only
observed difference in benthic community structure across the treatments was a slight decrease
in the abundance of Naididae oligochaetes at the highest spiking concentration. This absence
of any noteworthy biological response was consistent with the absence of interstitial water
concentrations of biological concern. This was attributed to the increase in concentrations of
iron and manganese sulfides, produced during periods of diagenisis, which were replaced by
the more stable zinc sulfide which is less readily oxidized during winter months. In this
experiment, and theoretically in nature, excesses of sediment metal might be overcome over
time due to the diagenisis of organic material. In periods of minimal diagenisis, the oxidation
rates of metal sulfides, if sufficiently great, could release biologically significant concentrations
of the metal into interstitial waters. This phenomenon should occur metal-by-metal in order of
their sulfide solubility product constants.
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52
SECTION 4
DERIVATION OF SEDIMENT GUIDELINES FOR METALS
4.1 GENERAL INFORMATION
Section 4 of this document presents the technical basis for establishing ESG for
copper, cadmium, nickel, lead, silver and zinc. The basis of the overall approach is the use of
EqP theory linked to the concept of maintaining metal activity for the sediment interstitial
water system below effects levels. Extensive toxicological data from short-term and long-term
laboratory and field experiments, with both marine and freshwater sediments and a variety of
species indicates that it is possible to reliably predict an absence of metal toxicity based upon
EqP theory. ESG for all six metals collectively can be derived using two procedures: (a) by
comparing the sum of their molar concentrations, measured as SEM, to the molar
concentration of AVS in sediments (AVS Guideline); or (b) by comparing the measured
interstitial water concentrations of the metals to WQC final chronic values (FCVs) (Interstitial
Water Guidelines). These approaches are described in more detail below. A lack of
exceedence of ESG based upon any one of the two procedures indicates that metal toxicity
should not occur. Exceedence of either the AVS or Interstitial Water Guidelines is indicative
of a potential problem that would entail further evaluation.
At present, EPA believes that the technical basis for implementing these two
approaches is supportable. The Organic Carbon and Minimum Partitioning Approaches as
proposed to the SAB and in Ankley et al.(1996) require additional research prior to their
implementation. Research issues for these latter two approaches include the development of
robust partitioning datasets for the six metals, as well as investigation of factors such as the
importance of other binding phases. The four approaches have been presented to and reviewed
by the Science Advisory Board of EPA (ILS EPA. 1994a; 1995a).
Additional research required to fully implement other approaches for deriving ESG
for these metals and to derive ESG for other metals includes the development of uncertainty
estimates associated with any approach; part of this would include their application to a variety
of field settings and sediment types. Research also is needed to establish the technical basis for
ESG for metals other than the six described herein, such as mercury, arsenic and chromium.
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53
Finally, the ESG approaches are intended to protect benthic organisms from direct toxicity
associated with exposure to metal-contaminated sediments. They are not designed to protect
aquatic systems from metal release associated, for example, with sediment suspension, or the
transport of metals into the food web either from sediment ingestion or the ingestion of
contaminated benthos. This latter issue, in particular, should be the focus of future research
given existing uncertainty in the prediction of bioaccumulation of metals by benthos (Ankley,
1996).
The following nomenclature is used in subsequent discussion of ESG derivation for
metals. The ESG for the metals are expressed in molar units because of the molar
stoichiometry of metal binding to AVS. Thus, solid phase constituents (AVS, SEM) are in
moles/g dry wt. The interstitial water metal concentrations are expressed in ^moles/L, either
as dissolved concentrations [Md] or activities {M2*} (Stumm and Morgan, 1981). The
subscripted notation, Md, is used to distinguish dissolved aqueous phase molar concentrations
from solid phase molar concentrations with no subscript. For the combined concentration,
[SEMT], the units are moles of metal per volume of solid plus liquid phase (i.e., bulk). Note
also that when [SEM^] is summed and/or compared, to AVS 1/2 the molar Ag concentration is
applied.
One final point should be made with respect to nomenclature. Use of the terms
non-toxic and having no effect, mean only with respect to the six metals considered in this
paper. The toxicity of field collected sediments can be caused by other chemicals. Therefore,
avoiding exceedences of ESG for metals does not mean that the sediments are non-toxic. It
only ensures that the six metals being considered should not have an undesirable biological
effect. Moreover, as discussed in detail below, exceedence of the guidelines for the six metals
does not necessarily indicate that metals will cause toxicity. For these reasons, we strongly
recommend the use of toxicity tests, TIEs, chemical monitoring in vertical, horizontal and
temporal scales, and other assessment methodologies as integral parts of any assessment
concerned with the effects of sediment-associated contaminants (Ankley et al., 1994).
4.2 SINGLE METAL SEDIMENT GUIDELINES
Except in rare instances, single metal guidelines are not usually applicable to field
situations since there is almost always more than one metal to be considered. As will become
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54
subsequently clear, it would be technically indefensible to derive guidelines for one metal at a
time because of the competitive nature of AVS binding. Nevertheless, it is illustrative to
present the logic for single metals as a prelude to the derivation of the multiple metal
guidelines.
4.2.1 AVS Guidelines
It has been demonstrated that if the SEM of a sediment is less than or equal to the
[SEM]* [AVS] (4-1)
AVS then no toxic effects are seen. This is consistent with the results of a chemical
equilibrium model for the sediment - interstitial water system (Di Toro et al., 1992). The
resulting metal activity {M2*} can be related to the total SEM of the sediment and water, and
to the solubility products of the metal sulfide (KMS) and iron sulfide (K^) . In particular, it is
true that at [SEM] & [AVS] then:
[SEMT]
Because the ratio of metal sulfide to iron sulfide solubility products (KMS/KFeS) is very small
(< 10"s) even for the most soluble of the sulfides, the metal activity of the sediment is at least
five orders of magnimde smaller than the SEM (see Di Toro et al. (1992) for data sources and
references): This indicates that no biological effects would be expected. Therefore, the
condition [SEM] < [AVS] is a "no effect" ESG.
The reason we use the term "no effect" is that for the condition [SEM] < [AVS] no
biological impacts are expected. However, for [SEM] > [AVS], which might seemingly be
considered a ESG violation, there are many documented instances where no biological impacts
occur (e.g., because organic carbon partitioning controls metal bioavailability in the interstitial
water, or the species of concern avoid or are insensitive to metals).
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55
4.2.2 Interstitial Water Guidelines
The condition [SEM] <; [AVS] indicates that the metal activity of the sediment -
interstitial water system is low and, therefore, below toxicologically-significant
concentrations. Another way of ensuring this is to place a condition on the interstitial water
activity directly. Suppose that we knew the metal activity, denoted by {FCV}, that
corresponded to the [FCV]. Then the ESG corresponding to this effect level is:
(4-3)
It is quite difficult, however, to measure and/or calculate metal activity in a solution phase, at
the low concentrations required, since it depends on the identities, concentrations and
thermodynamic affinities of other chemically reactive species that are present. Also the WQC
are not expressed on an activity basis. An approximation to this condition is:
[MdMFCVd]
(4-4)
where [FCVd] is the FCV applied to total dissolved metal concentrations. That is, we require
that the total dissolved metal concentration in the interstitial water [MJ be less than the FCV
applied as a dissolved guideline. Although this requirement ignores the effect of chemical
speciation on both sides of the equation - compare Equations (4-3) and (4-4) - it is the
approximation that is currently being suggested by EPA for the WQC for metals (Prothro,
1993). That is, the WQC should be applied to the total dissolved - rather than the total acid
recoverable - metal concentration (Table 4-1; U.S. EPA,l995b). Hence, if this second
condition is satisfied it is consistent with the level of protection afforded by the WQC.
In situations where the SEM exceeds the AVS ([SEMJ > [AVS]), but the interstitial
water total dissolved metal is less than the final chronic value ([MJ < [FCVJ), this sediment
would not violate the guidelines. These cases occur when significant binding to other phases
occurs. It should be noted that using the FCV for metals in freshwater samples requires that
the hardness of the interstitial water be measured since the WQC vary with1 hardness.
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56
4.3 MULTIPLE METALS GUIDELINES
As described in the previous subsection, from a practical standpoint it is insufficient
and inappropriate to consider each metal separately. Because of the interactive nature of
metal-sulfide binding, this is of particular concern for the AVS guidelines.
Table 4-1: Water quality criteria (WQC) criteria continuous concentrations (CCC) based on the
dissolved concentration of metallb. These WQC CCC values are for use in the Interstitial
Water Criteria Toxic Unit (IWCTU) approach for deriving sediment guidelines based on the
dissolved metal concentrations in interstitial water.
Metal
Cadmium
Copper*
Lead
Nickel
Silver
Zinc
Saltwater CCC, we/L
9.3
3.1
8.1
8.2
NAf
81
Freshwater CCC. uz/V
ppre(0.7852lln(h»rH-70S>]
0 997 [e<° W^WnOardneM)) +1.1 64S)i
NAf
0 986rei+0-76|4>i
(U.S.EPA, 1995b). .
Rounded all criteria to two significant figures.
For example the freshwater CCC at a hardness of 50, 100, and 200 mg CaCO3/L are 0.56,
0.94, and 1.6 Ag cadmium/L, 6.2, 12, and 20^g copper/L, 1.0, 2.5, and 6.1/zg lead/L, 88,
160, and 280^g nickel/L, and 58, 108, and 187 ng zinc/L.
CF= Conversion factor to calculate the dissolved CCC for cadmium from the total CCC for
cadmium: CF=1.101672-[(ln hardness)(0.041838)]
The saltwater CCC for copper is from the "Ambient Water Quality Criteria- Saltwater .
Copper Addendum" (U.S. EPA. I995c).
The silver criteria are currently under revision to reflect water quality factors that influence
the criteria such as hardness, and pH and any other factors. Since silver has the smallest
solubility product (see Table 2-2) and the greatest affinity for AVS, it would be the last
metal to be released from the AVS or the first metal to bind to the AVS so it is unlikely that
silver would occur in the interstitial water. However in sediments contaminated with silver
the user should be aware of the limitations in the above criteria for silver. AVS Guidelines
can be applied, however, the Interstitial Water Guidelines can not. If the AVS Guideline is
exceeded (£SEM > AVS) and the sediment is contaminated with silver, further testing and
evalutions would be warranted to access toxicity.
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57
4.3.1 AVS Guidelines
The results of calculations using chemical equilibrium models indicate that metals act in an
competitive manner when binding to AVS. That is, the six metals: silver, copper, lead,
cadmium, zinc and nickel will bind to AVS and be converted to their respective sulfides in
this sequence (i.e., in the order of increasing solubility). Therefore, they must be considered
together. There cannot be a guideline for just nickel, for example, since all the other metals
may be present as metal sulfides and, therefore, to some extent as AVS. If these other metals
are not measured as SEM, then the £SEM will be misleadingly small, and it may appear that
[£SEM] < [AVS] when in fact this would not be true if all the metals are considered together.
It should be noted that EPA currently restricts this discussion to the six metals listed above;
however, in situations where other sulfide forming metals (e.g., mercury) are present at high
concentrations, they also must be considered.
The equilibrium model prediction of the metal* activity is similar to the single metal example
when a mixture of the metals is present. If the molar sum of SEM for the six metals is less
than or equal to the AVS, that is:
[SEM^fAVS] ' (4.5)
then:
M K
(4-6)
[SEM,T] KFeS
where [SEMj>T] is the total SEM G"moles/L(bulk)) for the ilh metal. Thus the activity of each
metal, {Mt}, is unaffected by the presence of the other sulfides. This can be understood as
follows. Suppose that the chemical system starts initially as iron and metal sulfide solids and
that the system proceeds to equilibrium by each spl id dissolving to some extent. The iron
sulfide dissolves until the solubility product of iron sulfide is satisfied. This sets the sulfide
activity. Then each metal sulfide dissolves until reaching its solubility. Since so little of each
dissolve relative to the iron sulfide, the interstitial water chemistry is not appreciably changed.
Hence, the sulfide activity remains the same and the metal activity adjusts to meet each
solubility requirement. Therefore, each metal sulfide behaves independently of one another.
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58
The fact that they are only slightly soluble relative to iron sulfide is the cause of this behavior.
Thus, the AVS Guidelines are easily extended to the case of multiple metals.
4.3.2
Interstitial Water Guidelines
The application of the Interstitial Water Guideline to multiple metals is complicated, not by
the chemical interactions of the metals in the sediment - interstitial water system (as in the case
with the AVS Guideline), but rather because of their possible toxic interactions. Even if the
individual concentrations do not exceed the FCV of each metal (FCV,), the metals could exert
additive effects that might result in toxicity (Biesinger et al., 1986; Spehar and Fiandt, 1986;
Enserink et al., 1991; Kraak et al., 1994). Therefore, to address this potential additivity, the
interstitial water metal concentrations are converted to toxic units (TUs) and these are summed.
Since FCVs are used as the no effects concentrations these TUs are referred to as interstitial
water guidelines toxic units (IWGTUs). For freshwater sediments, the FCVs are hardness
dependent for all of the divalent metals under consideration and, thus, need to be adjusted to
the hardness of the interstitial water of the sediment being considered. Because there are no
FCVs for silver in freshwater or saltwater, this approach is not applicable to sediments
containing significant concentrations of silver (i.e., SSEM > AVS). Since silver has the
smallest solubility product (see Table 2-2) and the greatest affinity for AVS, it would be the
last metal to be released from the AVS or the first metal to bind with AVS so it is unlikely that
silver would occur in the interstitial water. For the i* metal with a total dissolved
concentration [Midl, the IWGTU is:
IWCTU;=
(4-7)
A lack of exceedence of the ESG requires that the sum of the IWGTUs be less than or equal to
one:
(4-8)
Hence, the multiple metals guideline is quite similar to the single metal case (Equation 4-4)
except that it is expressed as summed IWGTUs.
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59
To summarize, the proposed ESG are as follows. The sediment passes the ESG for the six
metals if either of these conditions is satisfied:
(a) AVS Guideline:
(4.5)
where
£ ilSEMJ = [SEMcJ + [SEMCd] + [SEMrt] + [SEMNj] -f [SEMZn] + [l/2SEMAg]
(b) Interstitial Water Guideline:
where
[MNid] [Mzn>d] [MAed]
[FCV.J [FCVCud] [FCVCdd] [FCVPM] [FCVNjd] [FCVZnd] [FCVAgd]
If either of these two conditions are violated, this does not mean that the sediment is toxic.
For example, if the AVS in a sediment is non-detectable, then condition (a) will be violated.
However, if there is sufficient organic carbon sorption so that condition (b) is satisfied, then
the sediment would be deemed acceptable.
If both of these conditions are violated, or if the AVS guideline is violated and the sediment
is contaminated with silver, then there is reason to believe that the sediment may be
unacceptably contaminated by these metals. Further testing and evaluations would therefore be
useful in order to assess actual toxicity and its causal relationship to the six metals. These may
include acute and chronic tests with species that are sensitive to the metals suspected to be
causing the toxicity. Also, in situ community assessments, sediment TIEs and seasonal
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60
characterizations of the SEM, AVS and interstitial water concentrations would be appropriate
(Ankleyetal., 1994).
4.4
ESG FOR METALS VS. ENVIRONMENTAL MONITORING DATABASES
The purpose of this Section is to compare ESG based on SEM-AVS or IWGTUs to chemical
monitoring data from freshwater and saltwater sediments in the United States. This
comparison of AVS-SEM and interstitial water concentrations can indicate the extent of metals
contamination in the United States. When toxicity or benthic organism community health data
are available in conjunction with these concentrations it is possible to speculate as to potential
causes of the observed effects.
4.4.1 Data Analysis
Three sources were identified which contain both AVS and SEM databases; one also had
data on concentrations of metals in interstitial water. Toxicity tests were also conducted on all
sediments from these sources. The databases are from the Environmental Monitoring and
Assessment Program (EMAP) (Leonard et al., 1996a), National Oceanic and Atmospheric
Administration, National Status and Trends Program (NOAA NS&T) (Wolf et al., 1994; Long
et al., 1995; 1996) and from the Regional Environmental Monitoring and Assessment Program
(REMAPHAdams et al., 1996).
Freshwater sediments:
The AVS and SEM concentrations in the 1994 EMAP database from the Great Lakes were
analyzed by Leonard et al. (1996a). Forty-six sediment grab samples and nine core samples
were collected in the summer from forty-two locations in Lake Michigan. SEM, AVS, TOC.
interstitial water metals (when sufficient volumes were present) and 10 day sediment toxicity to
the midge Chironomus tentans and the amphipod Hyallela aiteca were measured in sediments
collected by the grab (Appendix C).
The AVS concentrations vs. SEM-AVS differences from Appendix C are plotted in Figure
4-1. Grab sediment samples containing AVS concentrations below the detection limit of 0.05
umol/g AVS are plotted at that concentration. Forty-two of the 46 (91 percent) samples had
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10
E-10
a
CO
|-20
111
-30
•40
(a)
0.01
0.1 1 10
Acid Volatile Sulfide (pmo! S/g)
100
10
a
>
*
UJ
-5
(b)
-10
- 0.01
0.1 1 10
Acid Volatile Sulfide (pmol S/g)
100
Figure 4-1. SEM minus AVS values versus AVS concentrations in EMAP-Great Lakes
sediments from Lake Michigan. Data are from surficial grab samples only (this,figure is taken
from Leonard et al., 1996, see data in Appendix C). The upper plot shows all values, the
lower plot has the ordinate limited to SEM minus AVS values between -10 and +10.
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" ' 62.
SEM-AVS differences greater than 0. Thirty-six of these had less than 1.0 umol of £SEM
metal per gram sediment; and none had over 5.8 //moles/g of excess metal. In theory,
sediments with SEM concentrations in excess of that for AVS have the potential to be toxic due
to metals. However, the majority of exceedances occur in places where the AVS is very small
and the amount of SEM is also very small. For these Lake Michigan sediments, a closer look
at both interstitial water metal and toxicity test results is needed. Measurement of the
concentrations of metals in interstitial water can be used to determine if the excess metals are
bound to other sediment phases, therefore, prohibiting toxicity due to interstitial metal:
Interstitial water guidelines toxic units (IWGTU) can be calculated for each metal as the
interstitial water concentration divided by the final chronic value for that metal. Interstitial
water volumes were sufficient to measure metals concentrations in 20 of the samples. The sum
of the IWGTU for cadmium, copper, lead, nickel and zinc in these sediments was less than 0.4
(Leonard et al., 1996a). In 10-d toxicity tests using Chironomus teutons and Hyalella azteca,
no toxicity was observed 81 % of the 21 sediments not exceeding the ESG. They conclude that
for the toxic sediments that did not exceed the metals ESG, the observed toxicity is not likely
due to metals. Further, these sediments are unlikely to be contaminated by metals (Leonard et
al., 1996a). These data demonstrate the value of using both SEM-AVS and IWGTUs to
evaluate the risks of metals in sediments.
Saltwater sediments:
Saltwater data from a total of 398 sediment samples from five monitoring programs
representing the eastern coast of the United States from Chesapeake Bay to Massachusetts are
included in Figure 4^2. The EMAP Virginia Province database (U.S. EPA, 1996) consists, in
part, of 127 sediment samples collected from August to mid-September 1993 from randomly
selected locations in tidal rivers and small and large estuaries from the Chesapeake Bay to
Massachusetts (Strobel et al., 1995). The NOAA data is from Long Island Sound, Boston
Harbor and the Hudson River Estuary. Sediments were collected from 63 locations in the
coastal bays and harbors of the Long Island Sound in August, 1991 (Wolfe et al., 1994).
Sediment samples from 30 locations in Boston Harbor were collected in June and July 1993
(Long et al., 1996). Sediment samples from 38 locations in the Hudson River Estuary were
collected from March to May 1991 (Long et al., 1995). Sediment samples were collected in
the REMAP program from 140 locations from the New York/ New Jersey Harbor Estuary
System (Adams et al., 1996). All of the above sediment grab samples were from
approximately the top 2 cm of undisturbed sediment.
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64
For saltwater sediments, molar concentration of AVS typically exceeds that for SEM (SEM-
AVS <0) for most of the samples across the entire range of AVS concentrations (Figure 4-2).
A total of 68 of the 398 saltwater sediments (17 percent) had an excess of metal, and only 4 of
the 68 (6 percent) had over 2 umol/g excess SEM. As AVS levels increase above this
concentration fewer and fewer sediments have SEM-AVS differences that are positive; none
occurred when AVS was > 8.1 umol/g. Unlike the sediments from the freshwater EMAP
survey in Lake Michigan, interstitial .water was not measured in these saltwater sediments.
Only five of the 68 sediments (7 percent) having excess of up to 0.9 umol/g SEM were toxic in
10-d sediment toxicity tests with the amphipod Ampelisca abdita, whereas 79 of 330 (24
percent) sediments having an excess of AVS were toxic. The data support the interpretation
that (1) toxicity was NOT metals-related in the 79 sediments where AVS was in excess over
SEM; (2) metals might have caused the toxicity in the five toxic sediments having an excess of
metal, but even in the absence of measurements of interstitial water metals concentrations, we
speculate that metals toxicity is unlikely because there was only $0.9 umol/g excess SEM (the
molar concentration SEM most often exceeds that of AVS, in sediments having AVS
concentrations s 1 umol/g); and (3) the absence of toxicity in sediments having an excess of
SEM of up to 4.4 umol/g indicates that significant metal-binding potential over that of AVS
existed in some sediments. Organic carbon concentrations of from 0.05% to 15.2% (average
1.9 percent) provides for some of this additional metal-binding.
The data above appear to suggest that in the United States direct toxicity caused by metals in
sediments is extremely rare. While this might be true,these data by themselves are
inconclusive and it would be inappropriate to use the data from the above studies to reach this
conclusion. All of the above studies were conducted in the summer when the seasonal
biogeochemical cycling of sulfur should produce the highest concentrations of iron monosulfide
which should make direct metal-associated toxicity less likely than in the winter/spring months.
Accurate assessment of the extent of the direct ecological risks of metals in sediments requires
that sediment monitoring occur in the months of minimum AVS concentration; typically
November to early May. These yet to be conducted studies must monitor at a minimum SEM,
AVS, interstitial water metal and toxicity. The data presented here are not intended to be used
to draw conclusions about toxicity due to resuspension or bioaccumulation.'
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SECTION 5
IMPLEMENTATION
5.1 CONSIDERATIONS IN PREDICTING METAL TOXICITY
Results of the short- and long-term laboratory and field experiments conducted to date
using sediments spiked with individual metals and mixtures of metals represent convincing
support of the conclusion that an absence (but not necessarily a presence) of metal toxicity can
be reliably predicted based upon metal:sulfide relationships and/or interstitial water metal
concentrations. In contrast, much confusion exists in the use of this convincing evidence to
interpret the significance of metals concentrations in sediments from the field when toxicity
and benthic community structure measurements are available. In addition, the use of these
observations as a basis for predicting metal bioavailability, or deriving ESG, raises a number
of conceptual and practical issues related to sampling, analytical measurements and effects of
additional binding phases. Many of these were addressed by Ankley et at. (1994); the most
salient to the proposed derivation of ESG are described below.
5.2 SAMPLING AND STORAGE
Accurate prediction of exposure of benthic organisms to metals is critically dependent
upon sampling appropriate sediment horizons at appropriate times. This is because of the
relatively high rates of AVS oxidation due to natural processes in sediments and the
requirement that oxidation must be avoided during sampling of sediments and interstitial water.
In fact it is this seemingly labile nature that has led some to question the practical utility of
using AVS as a basis for EqP-derived ESG for metals (Luorria and Carter, 1993; Meyer et al..
1994). For example, there have been many observations of spatial (depth) variations in AVS
concentrations, most of which indicate that surficiat AVS concentrations are less than those in
deeper sediments (Besser et al., 1996; Boothman and Helmstetter,1992; Brumbaugh et al.,
1994; Hansen et al., 1996b; Hare et al., 1994; Howard and Evans, 1993; Leonard et al.,
1996a; Liber et al., 1996 ). This likely is due to oxidation of AVS at the sediment surface, a
process that is enhanced by bioturbation (Peterson et al., 1996). In addition to varying with
depth, AVS can vary seasonally. For example, in systems where overlying water contains
appreciable oxygen during cold weather months, AVS tends to decrease, presumably due to a
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constant rate of oxidation of the AVS linked to a decrease in its generation by sulfate-reducing
bacteria (Herlihy and Mills, 1985; Howard and Evans, 1993; Leonard et al., 1993). Because
of potential temporal and spatial variability of AVS, it appears the way to avoid possible
wider-estimation of metal bioavailability is to sample the biologically "active" zone of
sediments at times when AVS might be expected to be present at small concentrations. We
recommend that at a minimum AVS and SEM measurements be made using surficial (0-2.0
cm) sediments during the period from November to early May in aerobic aquatic ecosystems. .
Minimum AVS concentrations may not always occur during cool-weather seasons; for
example, systems that become anaerobic during the winter can maintain relatively large
sediment AVS concentrations (Liber et al., 1996). Therefore, seasonal measurements of AVS,
SEM and interstitial metal concentrations may need to be determined. Importantly, the •
biologically active zones of some benthic communities may be only the surface few millimeters
in depth and in other communities up to a meter. Therefore, for sufficient characterization,
multiple sediment horizons may require sampling of interstitial water, SEM and AVS to
determine the potential for exposure to metals.
The somewhat subjective aspects of these sampling recommendations have been of
concern. However, recent research suggests that the transient nature of AVS may be over-
stated relative to predicting the fate of all metal-sulfide complexes in aquatic sediments.
Observations from the Duluth EPA laboratory made in the early 1990s indicated that AVS
concentrations in sediments contaminated by metals such as cadmium and zinc tended to be
elevated over concentrations typically expected in freshwater systems (G.T. Ankley,
unpublished data). The probable underlying basis for these observations did not become
apparent, however, until a recent series of spiking and metal-sulfide stability experiments. The
field colonization study of Liber et al. (1996) demonstrated a strong positive correlation
between the amount of zinc added to test sediments and the resultant concentration of AVS in
the samples. In fact, the initial design of their study attempted to produce test sediments with
nominally as much as five-times more SEM (zinc) than AVS; however, the highest measured
SEM/AVS ratio achieved was only slightly larger than 1. Moreover, the expected surficial
depletion and seasonal variations in AVS'were unexpectedly low in the zinc-spiked sediments.
These observations suggested that zinc sulfide, which comprised the bulk of AVS in the spiked
sediments, was more stable than the iron sulfide that presumably was the source of most of the
AVS in the control sediments. The apparent stability of other metal sulfides versus iron sulfide
also has been noted in laboratory spiking experiments with freshwater and saltwater sediments
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(Boothmanet al., 1996; De Witt et al., 1996; Hansenet a!., 1996b; Leonard et al., 1995;
Peterson et al,, 1996; Sibley et al., 1996).
In support of these observations, recent metal-sulfide oxidation experiments conducted
by Di Toro et al. (1996b) have confirmed that cadmium and zinc form more stable sulfide solid
phases than iron. If this is also true for sulfide complexes of copper, nickel and lead, the issue
of seasonal/spatial variations in AVS becomes of less concern because most of the studies
evaluating variations in AVS have focused on iron sulfide (i.e., uncontaminated sediments).
Thus, further research concerning the differential stability of metal-sulfides, both from a
temporal and spatial perspective, is definitely warranted.
5.2.1 Sediments
At a minimum, sampling of the surficial 2.0 cm of sediment in between November and
early May is recommended. A sample depth of 2.0 cm is more appropriate for remediation
and monitoring. In some instances such as for dredging or where depths greater than 2 cm are
important than sample depths should.be planned based on particular study needs. Sediments
. can be sampled using dredges, grabs, or coring, but mixing of aerobic and anaerobic sediments
must be avoided because the trace metal speciation in the sediments will be altered (See Bufflap
and Allen, 1995 for detailed recommendations to limit sampling artifacts). Coring is generally
* less disruptive, facilitates sampling of sediment horizons and limits potential metal
contamination and oxidation if sealed PVC core liners are used.
Sediments not immediately analyzed for AVS and SEM must be placed in sealed air-
tight glass jars and refrigerated or frozen. Generally, 50 ml or more of sediment should be
added to nearly fill the jar. If sediments are stored this way there will be little oxidation of
AVS even after several weeks. Sampling of the stored sediment from the middle of the jar will
further limit potential effects of oxidation on AVS. Sediments experiencing oxidation of AVS
during storage will become less black or grey if oxidized. Because the rate of metal-sulfide
oxidation is markedly less than that of iron sulfide, release of metal during storage is unlikely.
5.2.2 Interstitial Water
Several procedures are available to sample interstitial water in situ or ex situ. Carignan
et al. (1985) compared metals concentrations in interstitial water .obtained by ex situ
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centrifugation at 11,000 rpm followed by double filtration (0.45 /^m and 0.2 or 0.03 #m) and
in situ diffusion samplers with a 0.002 /urn interstitial size. For the metals of concern in this
guidelines document, concentrations of nickel and cadmium were equivalent using both
methods and concentrations of copper and zinc were higher and more variable using
centrifugation. They recommended the use of in situ dialysis for the study of trace constituents
in sediments because of its inherent simplicity and the avoidance of artifacts that can occur
with the handling of sediments in the laboratory.
• r
More recently Bufflap and Allen (1995) reviewed four procedures for the collection of
interstitial water for trace metals analysis. Thesejncluded ex situ squeezing and cenfrifiigation^
andjnjitu dialysis and suction filtration. This paper should be read by thosejejecting a
interstitial water sampling method. They observed that each method has its own advantages
and disadvantages, and that each user must make their own choice given the inherent errors of
each method. Importantly, interstitial water must be extracted by centrifugation or squeezing
in an inert atmosphere until acidified because oxidation will alter metal spectation. Artifacts
may be caused by temperature changes in ex situ methods that may be overcome by
maintaining temperatures to those in situ. Contamination of interstitial water by fine particles is
important in all methods as differentiation of paniculate and dissolved metal is a function of
interstitial size. The use of 0.45 ton filtration, while an often accepted definition of dissolved,
may result in laboratory to laboratory discrepancies. The use of suction filtration devices is
limited to coarser sediments, and they do not offer depth resolution. The use of diffusion
samplers is hampered by the time required for equilibrium (7-14 days) and the need for diver
placement and retrieval in deep waters. Acidification of interstitial water obtained by
diffusion or from suction filtration must occur immediately to limit oxidation. Bufflap and
Allen (1995) conclude that in situ techniques have less potential for producing sampling
artifacts than ex situ procedures. They concluded that of the in situ procedures, suction
filtration has the best potential for producing artifact free interstitial water samples directly
from the environment. Of the ex-situ procedures they concluded that centrifugation under a
nitrogen atmosphere followed immediately by filtration and acidification was the simplest
technique likely to result in an unbiased estimate of metal concentrations in interstitial water.
At present, EPA recommends filtration of the surface water through 0.4 to 0.45 y.
polycarbonate filters to better define that fraction of aqueous metal associated with toxicity
(Prothro, 1993). Thurman equates the organic carbon retained on a 0.45 micrometer glass-
fiber filter to suspended organic carbon so that this filtration procedure under nitrogen
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atmosphere followed immediately by acidification is acceptable for interstitial waters.
However in studies comparing collection and processing methods for trace metals, sorption to
filter membranes or filtering apparatus has been identified when losses occur (Ozretich and
Schults, 1998). Ozretich and Schults, 1998 have recently presented a method combining
longer centrifiigation times with a unique single-step IW withdrawal procedure which has the
potential for minimizing metal losses by eliminating the need for filtration.
In contrast to the above recommendations, EPA recommends the use of dialysis
samplers to obtain samples of interstitial water for comparison of measured concentrations of
dissolved metals with WQC. This is primarily because diffusion samplers obtain interstitial
water with the proper in situ geochemistry thus limiting artifacts of ex situ sampling. Further,
EPA has found that in shallow waters where contamination of sediments is most likely,
placement of diffusion samplers is easily accomplished and extended equilibration times are not
a problem. Secondly, EPA recommends the use of centrifiigation under nitrogen and double
0.45/um filtration using polycarbonate filters for obtaining interstitial water from sediments in
deeper aquatic systems. Probably most importantly, the extremely large database comparing
interstitial metals concentrations with organism responses from spiked and field sediment
experiments in the laboratory has demonstrated that, where the interstitial water toxic unit
concept predicted that metals concentrations in interstitial water should not be toxic, toxicity
was not observed when either dialysis samplers or centrifugation were used (Berry et al.,
1996a; Hansen et al., 1996a). Therefore, it is likely that when either methodology is used to
obtain interstitial water for comparison with WQC, if metals concentrations are below 1.0
IWGTU sediments should be acceptable for protection of benthic organisms.
5.3 ANALYTICAL MEASUREMENTS
An important aspect to deriving "global" ESG values is that the methods necessary to
implement the approach must be reasonably standardized or have been demonstrated to
produce results that are comparable to those of standard methodologies. From the standpoint
of the proposed metal ESG, a significant amount of research has gone into defining
methodologies to obtain interstitial water and sediments (see Section 5-2 above), to extract
SEM and AVS from sediments, and to quantify AVS, SEM and the metals in interstitial water.
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5.3.1 Acid Volatile Sulfide
The SEM/AVS extraction method recommended by EPA is that of Allen et al. (1993).
In terms of AVS quantification, a number of techniques have been successfully utilized
including gravimetric (Di Toro et al., 1990; Leonard et al., 1993), colorimetric (Cornwell and
Morse, 1987), gas chromatography - photoionization detection (Casas and Crecelius, 1994;
Slotton and Reuter, 1995) and specific ion electrodes (Boothman and Helmstetter,1992;
Brouwer and Murphy, 1994; Brumbaugh et al., 1994; Leonard et al., 1996b). Allen et al.,
1993 report a limit of detection for 50% accuracy of 0.01 umol/g for a 10-g sediment sample
using the colorimetric method. Based on several studies Boothman reports a detection limit of
1 umol AVS which translates to 0.1 umol/g dry weight for a 10 g sediment sample using the
ion specific electrode method (personal communication).
5.3.2 Simultaneously Extracted Metal
Simultaneously extracted metals are operationally-defined as metals extracted from
sediment into solution by the acid volatile sulfide extraction procedure. The "dissolved"
metals in this solution are also operationally defined as the metal species which pass through
filter material used to remove the residual sediment, and thus are defined by the interstitial size
of the filtration material used. Common convention defines "dissolved" as metal species
<0.45-#m.in size. SEM concentrations measured in sediments are not significantly different,
however, using Whatman 1 filter paper alone (< 11-jtm nominal interstitial size) or in
combination with a 0.45-/im filter (W. Boothman, unpublished data). SEM solutions generated
by the AVS procedure can be analyzed for metals, commonly including cadmium, copper,
lead, nickel, silver and zinc by routine atomic spectrochemical techniques appropriate for
environmental waters (e.g. inductively coupled plasma atomic emission or graphite furnace
atomic absorption spectrophotometry) (U.S. EPA, I994b). Because of the need to determine
metals at relatively low concentrations, additional consideration must be given to preclude
contamination during collection, transport and analysis (U.S. EPA, 1995d,e,f,g).
5.3.3 Interstitial Water Metal
Interstitial water can be analyzed for the metals cadmium, copper, lead, nickel, silver
and zinc by routine atomic spectrochemical techniques appropriate for environmental waters
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(e.g. inductively coupled plasma atomic emission or graphite furnace atomic absorption
spectrophotometry) (U.S. EPA, 1994b). Because of the need to determine metals at
concentrations at or below the threshold of biological effects (i.e., WQC concentrations);
additional consideration must be given to preclude contamination during collection, transport
and analysis (U.S. EPA, 1995d,e,f,g). (See guidance on clean chemistry techniques in U.S.
EPA, 1994c.) Generally, detection limits should be at sO.l IWGTU, because the toxic unit
contributions of each of the metals must be summed.
5.4 ADDITIONAL BINDING PHASES
Although AVS is an important binding phase for metals, there clearly are other
physico-chemical factors that influence metal partitioning in sediments. In aerobic systems, or
those with low productivity (i.e., where the absence of organic carbon limits sulfate reduction),
AVS plays little or no role in determining interstitial water concentrations of metals. For
example, Leonard et al. (1996a) found that a relatively large percentage of surficial sediments
from open areas in Lake Michigan did hot contain detectable AVS. In fact the great majority
(42 of 46) of samples analyzed by Leonard et al. (1996a) contained less AVS than SEM, yet
interstitial water metal concentrations of cadmium, copper, nickel, lead and zinc were
consistently small or non-detectable. Even in sediments where concentrations of AVS are
significant, other partitioning phases may provide additional binding capacity for SEM (e.g.,
Ankley et 41., 1993; Calamono et al., 1990; Slotton and Reuter, 1995). In aerobic sediments
both organic carbon and iron and manganese oxides control interstitial water concentrations of
metals (Calamono et al,. 1990; Jenne, 1968; Luoma and Bryan, 1981; Tessier et al., 1979). In
anaerobic sediments, organic carbon appears to be an important additional binding phase
controlling metal partitioning, in particular for cadmium, copper and lead (U.S. EPA, 1994a).
Even in substrates with very little metal binding capacity, (e.g., chromatographic sand).
surface adsorption associated with caiion exchange capacity will control interstitial water metal
concentrations to some degree (Hassan et al., 1996). Although an ideal ESG model for metals
would incorporate all possible metal binding phases, current knowledge concerning
partitioning/capacity of phases other than AVS is insufficient for practical application of a
multiple phase model for deriving ESG in this sediment guidelines document.
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5.5 PREDICTION OF THE RISKS OF METALS IN SEDIMENTS BASED ON EqP
It is important to repeat that conclusions about sediment toxicity based on SEM-AVS
concentrations pertain only to cadmium, copper, lead, nickel, silver and zinc. (1) When the
molar concentration of AVS exceeds that of SEM (negative SEM-AVS) sediment toxicity due
to these metals is unlikely and any observed toxicity is most likely from some other cause. This
is important because toxicity observed in sediments having an excess of AVS is often
incorrectly assumed to disprove the EqP metals theory. The correct conclusion is that some
factor other than metals caused the effect. This can be further substantiated if the toxic unit
concept is applied to metal concentrations measured in interstitial water; the absence of
significant concentrations of metals coupled with the negative SEM-AVS are powerful
evidence that metals are an unlikely cause of the effect. (2) Sediments can only be toxic from
the metals cadmium, copper, lead, nickel, silver and zinc when the molar concentrations of
SEM exceed those of AVS (SEM-AVS differences are positive). Measurements of interstitial
water concentrations of metals are invaluable in demonstrating that the sediments are toxic
because of metals, and these measurements will provide insights into the specific metal(s)
causing the observed toxicity. (3) It is not uncommon for toxicity to be absent in sediments
having concentrations of SEM that exceed those of AVS (SEM-AVS is positive). This is
because other metal binding phases in sediments often reduce the concentrations of bioavailable
metal. (4) When sediments are toxic, and SEM-AVS is greater than 0.0, the toxicity may or
may not be metals-related. Often sediments having SEM-AVS of up to 10 //moles SEM/g are
not toxic because the excess metals are associated with other binding phases. Measurements of
interstitial water concentrations of metals are invaluable in demonstrating an absence or
presence of bioavailable metal.
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SECTION 6
GUIDELINES STATEMENT
The procedures described in this document indicate that, except possibly where a
locally important species is very sensitive, benthic organisms should be acceptably protected in
freshwater and saltwater sediments if any one or both of the following two conditions is
satisfied: (a) If the sum of the molar concentrations of SEM cadmium, copper, lead, nickel,
silver and zinc is less than or equal to the molar concentration of AVS or (b) the sum of the
dissolved interstitial water concentration of cadmium, copper, lead, nickel, silver and zinc
divided by their respective WQC is less than or equal to 1.0.
(a) AVS Guidelines:
, [SEMJslAVS] (4-5)
where
£ ilSEMJ = [SEMCa] + [SEMCd] + [SEMpJ + [SEMNl] + [SEMZn] + [1/2SEM*,]
(b) Interstitial Water Guidelines
where
[MNidj ^ [M2nd] [MAsd]
[FCV,dl [FCVCBd]cVC(td] [FCVpbd] [FCVNJd] [FCVZ J [FCVAgd]
If any one of these two conditions are violated, this does not mean that the sediment violates the
ESG and is unacceptable. For example, if SEM exceeds AVS, or if the AVS in a sediment is non-
detectable, then condition (a) will be violated. However, if there is sufficient sorption to particles,
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or organic carbon or other binding phases so that condition (b) is satisfied, then the sediment meets
the guideline and benthic organisms are acceptably protected from metals-induced sediment
toxicity.
If both of these conditions are violated, or if the AVS Guideline is violated and the
sediment is contaminated with silver then there is reason to believe that the sediment may be
unacceptably contaminated by these metals. Further testing and evaluations would, therefore, be
useful in order to assess actual toxicity and its causal relationship to the five metals. These may
include acute and chronic tests with species that are sensitive to the metals suspected to be causing
the toxicity. Also, in situ community assessments, sediment TIEs and seasonal characterizations
of the SEM, AVS and interstitial water concentrations would be appropriate (Ankley et al., 1994).
The ESG approaches are intended to protect benthic organisms from direct toxicity
associated with exposure to metal-contaminated sediments. They are not designed to protect
aquatic systems from metal release associated, for example, with sediment suspension, or the
transport of metals into the food web either from sediment ingestion or the ingestion of
contaminated benthos. This latter issue, in particular, should be the focus of future research given
existing uncertainty in the prediction of bioaccumulation of metals by benthos (Ankley, 1996).
It is repeated here that these guidelines apply only to the six metals discussed in this
document, copper cadmium; lead, nickel, zinc and silver. Procedures for sampling and analytical
methods for interstitial water and sediments are discussed in Section 5, Implementation.
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SECTION 7
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APPENDIX A
-------
ACR
Ag
AVS
ASTM
Cd
ccc
CFR
CLOGP
CMC
CV
CWA
DOC
EDTA
EMAP
ESG
foe
EqP
FAV
FCV
fFeS(s)]
[FeS(s)],
FeS
FPV
FRY
GC/EC
GC/MS
APPENDIX A
Glossary of Abbreviations and Equations
Acute-Chronic Ratio
Silver
Acid Volatile Sulfide
American Society for Testing and Materials
Cadmium
Freely dissolved interstitial water concentration of contaminant
Total interstitial water concentration of contaminant
Concentration of contaminant in sediment
Concentration of contaminant in sediment on an organic carbon basis
Criteria Continuous Concentration
Code of Federal Regulations
Computer program for generating partition coefficients
Criteria Maximum Concentration
Coefficient of Variation :
Clean Water Act ' "
Dissolved Organic Carbon
Ethlyene diamine tetraacetic acid
Environmental Monitoring and Assessment Program
Equilibrium Partitioning Sediment Guidelines
. Organic carbon-normalized Equilibrium Partitioning Sediment Guidelines
Fraction of organic carbon in sediment
Equilibrium partitioning
Final Acute Value
Final Chronic Value
activity of Fe:* (mol/L)
concentration of Fe2" (mol/L)
concentration of iron sulfide (mol/L),
initial iron sulfide concentration in the sediment (mol/L)
Iron monosulfide
Final Plant Value • '
Final Residue Value
Gas Chromatography/Electron Capture
Gas Chromatography/Mass Spectrometry
-------
GFAA
IWGTU
IWTU
KP
KSP
LC50
L
[M2+]
[M]A
[MS(s)J
m3 or cu m
jumole
mg
mg/1
ml
mm
NA
ND
ng
Nl
NOAA
NOEC
NST
NTA
Pb
pH
DEC
POC
Graphite Furnace Atomic Absorption
Interstitial Water Guidelines Toxic Unit
Interstitial Water Toxic Unit
solubility product for FeS(s) [(mol/L)2]
solubility product for MS(s) [(mol/L)2]
Organic carbon: water partition coefficient
Octanol: water partition coefficient
Sediment: water partition coefficient
Solubility product constant
Concentration estimated to be lethal to 50 percent of the test organisms within a
specified time period.
Liter
divalent metal activity (mol/L)
concentration of M2* (mol/L)
concentraton of added metal (mol/L)
concentration of solid-phase metal sulfide (mol/L)
Cubic meter
Microgram •
Micrometer
Micromole
Milligram
Milligram per liter
Milliliter
Millimeter
Not Applicable, Not Available
Not Determined, Not Detected
Nanogram .
Nickel
National Oceanographic and Atmospheric Administration
No Observed Effect Concentration
National Status and Trends monitoring program
Nitrilotriacetic acid
Lead
Negative logarithm of the effective hydrogen ion concentration
Observed Effect Concentration
Paniculate Organic Carbon
-------
APPENDIX B
-------
ppb
ppm
ppt
REMAP
{S2-}
[S2']
[SEM]
lSEM]Cd
[SEM]Cu
[SEM]Ni
[SEMlp,
SAB
SD
SLC
SEM
SOP
STORET
TDS
TOC
TU
TVS
U.S. EPA
WQC
Zn
Ys?-
[2Fe(aq)]
[SM(aq>]
[ZS(aq)]
Parts per billion
Parts per million
Pans per trillion
Regional Environmental Monitoring and Assessment Program
activity of S2' (mol/L)
concentration of S2" (mol/L)
simultaneously extracted metal concentration (/imol/g)
simultaneously extracted Cd concentration (/xmol/g)
simultaneously extracted Cu concentration 0*mol/g)
simultaneously extracted Ni concentration (/xmol/g)
simultaneously extracted Pb concentration (jtmol/g)
simultaneously extracted Zn concentration (/xmol/g)
U.S. EPA Science Advisory Board
Standard Deviation
Screening Level Concentration
Simultaneously Extracted Metals
Standard Operating Procedure
EPA's computerized water quality data base
Total Dissolved Solids
Total Organic Carbon
Toxic Unit
Total Volatile Solids
United States Environmental Protection Agency
Water Quality Criteria
Zinc
{Fe2+}/[2Fe(aq)]
+ }/[2m(aq)]
activity coefficient of Fe'+
activity coefficient of M'+
activity coefficient of S2"
concentration of total dissolved Fe(II) (mol/L)
concentration of total dissolved M(II) (mol/L)
concentration of total dissolved S(II) (mol/L)
-------
asJ. = [S2l/[£S(aq)]
(B-8)
are the ratios of the divalent species concentrations to the total dissolved M(II), Fe(II), and S(-
H) concentrations, [SM(aq)], [SFe(aq>], and [2S(aq)J, respectively. [MS(s)] and [FeS(s>] are
the concentrations of solid-phase metal and iron sulfides at equilibrium. [FeS(s)]j is the initial
iron sulfide concentration in the sediment, and [M]A is the concentration of added metal.
The solution of these five equations can be obtained as follows. The mass balance
Equations B-3 and B-4 for M(II) and FE(H) can be solved for [MS(s)J and [FeS(s>] and
substituted in the mass balance Equation B-5 for S(II):
a'VfFe2*] + cc'VlM2*] = [Ml
(B-9)
The mass action Equations B-l and B-2 can be used to substitute for [Fe2*] and [M2*], which
results in a quadratic equation for [S2~]:
-a
YS2-[S2-]
VM-
= [MJ,
(B-10)
The positive root can be accurately approximated by:
[ML
(B-ll)
which results from ignoring the leading term in Equation B-10. This is legitimate because the
-------
APPENDIX B
Solubility Relationships for Metal Sulfides
Consider the following situation: a quantity of FeS is titrated with a metal that forms a
more insoluble sulfide. We analyze the result using an equilibrium model of the M-(n)-Fe(D)-
S(-II) system. The mass action laws for the metal and iron sulfides are
(B-2)
where [M2+], [Fe2*] and [S2'] are the molar concentrations; YM2.,yfti. andJ^j. are the activity
coefficients; and KMS and KFeS are the sulfide solubility products. The mass balance equations
for total M(H), Fe(II) and S(-II) are
(B-3)
'*] + [FeS(s)j = [FeS(s)].
a" S:-[SJ'3 * [MS(s)] - [FeS(s)] = [FeS(s)]. (B-5)
where
«M:-'- [M:-]/[EM(aq)] . (B.6)
aFe:. = [Fe2']/[rFe(aq)] ' - (B.7)
-------
{M2*}
(B-16)
The magnitude of the term in parentheses can be estimated as follows. The first term in the
denominator is always greater than or equal to 1, PFe2+J> I, because it is the reciprocal of two
terms both of which are less than or equal to 1, Equation B-14. They are aFe2*_< 1, which is
the ratio of the divalent to total aqueous concentration, and vFe2+ _< 1, which is an activity
coefficient. The second term in the denominator cannot be negative, PM2*KMS/KFeS > 0, since
all of its terms are positive. Thus, the denominator of the expression in parentheses is always
greater than 1, pFe2+ + PM2*KMS/KFeS > 1. Therefore, the expression in parentheses is always
less than 1. Hence, the magnitude of the ratio of metal activity to total added metal is bounded
from above by ratio of the sulfide solubility products:
{Me2*}/[M],
KMS/KFeS
(B-17)
This results applies if [FeS]j > [M]A so that excess [FeS(s)] is present.
If sufficient metal is added to exhaust the initial quantity of iron sulfide, then [FeS(s)]
= 0. Hence, the iron sulfide mass action equation (B-2) is invalid and the above equation no
longer applies. Instead, the only solid-phase sulfide is metal sulfide and
[MS] -~ [FeS],
(B-18)
so that, from the metal mass balance equation
^> = Y,/W([M]A - [FeS(s).)
(B-19)
this completes the derivation of Equations 2-8 and 2-9.
-------
term in parentheses in Equation B-10 is small relative to [M]A due to the presence of the sulfide
solubility products. As a result, [S2*] is also small since it is in the denominator. Hence, the
leading term in Equation B-10 must be small relative to [M]A and can safely be ignored.
The metal activity can now be found from the solubility equilibrium Equation B-l:
{M2*} =
1 ' KMS
Yc2-[
-i
[M]
so that
(B-13)
where
and
(B-14)
(B-15)
Equation A-13 can be expressed as
-------
APPENDIX C
^Boncentratiqns of SEM, AVS, TOC, and IWCTU for cadmium, copper, lead, nickel, and zinc in 46 surficial samples from Lake Michigan
Sam- TOC
pie (%)
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
M5
p*
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
0.18
4.63
3.36
4.89
0.92
4.37
5.27
0.08
4.27
2.11
1.89
0.41
2.87
3.68
0.28
0.07
3.51
0.40
1.73
0.69
2.51
1.17
0.13
1.03
0.63
0.30
6.29
0.21
0.11
0.05
0.27
4.95
0.54
6.75
0.18
0.15
0.56
0.10
0.06
2.68
0.16
1.80
1.29
0.05
0.14
0.57
SEM
(umol/g)
0.53
3.46
2.78
3.55
0.14
2.82
1.20
0.17
1.47
0.25
1.12
0.74
1.17
1.56
1.32
0.17
0.75
0.97-
1.74
0.70
0.19
0.59
0.21
0.62
0.13
0.15
0.25
0.12
0.20'
0.04
0.85
1.17
' 0.44
1.37
0.26
0.06
0.17
0.22
0.06
5.83
0.16
0.56
1.02
0.06
0.16
0.66
AVS SEM-
(umol/g) AVS
0.03
0.35
0.06
0.05
0.03
1.13
0.13
0.03
4.49
0.03
0.03
0.07
0.18
0.03
0.44
0.05
0.08
0.03
0.15
0.03
0.05
0.03
0.03
0.03
0.20
0.03
0.03
0.03
0.06
0.03
0.03
1.66
0.12
0.09
0.03
0.05
0.05
0.12
0.03
0.03
0.07
0.03
2.25
0.03
0.05
0.03
0.51
3.11
2.72
3.50
0.12
1.69
1.07
0.15
-3.02
0.23
1.10
0.67
0.99
1.54
0.88
0.12
0.67
0.95
1.59
0.68
0.14
0,57
. 0.19
0.60
-0.07
0.13
0.23
0.10
0.14
0.02
" 0,83
-0.49
0.32
1.28
0.24
0.01
0.12
0.10
0.04
5.81
' 0.09
0.54
-1.23
0.04
0.11
0.64
Cadmium
0.029
0.0(8
0.018
0.0002
0.024
0.029
0.115
0.050
'
-
0.0002
.
0.0002
0.0002
-
0.018
-
0.079
.
-
-
- .
-
.
'
* -
0.0002
0.0002
'
.
0.012
-
0.018
-
«
.
• ' -
-
; 0.003
'
0.006
0-0002
-
.
-
Copper
..
0.003
0.308
0.266
0.034
0.049
0.003
0.003
0.034
.
.
0.070
.
0.003
0.119
-
0.060
-
0.013
.
.
-
.
.
.
.
.
0.155
0.003
-
-
0.036 •
-
0.041
-
-
-
-
-
0.119
' -
0.003
0.028
-
-
-
Lead
0.00004
0.002
0.0004
0.0008
0.0002
0.0001
0.001
0.0008
-
-
0.002
.
0.0004
0.0002
-
0.0008
-
0.0008
.
.
-
.
. .
.
.
.
0.0001
0.0004
- •
.
0.0004 "
-
0.0002
-
'.
.
.
- •
0.001
-
0.0006
0.002
.
.
•
IWCTU
Nickel
.
0.005
0.003
0.003
0.006
0.004
0.006
0.006
0.004
-
-
0.0005
.
0.006
0.004
-
0.008
- .
0.010
.
-
-
.
• .
.
.
-
0.011
0.007
-
.
0.002
-
0.017
-
.,
. •_
.
-
0.0005
-
0.008
0.0005
-
.
-
% Survival
Zinc
.
0.003
0.029
0.006
0.032
0.020
0.020
0.055
0.026
-
-
0.001
.
0.015
0.050
-
0.058
-
0.020
.
-
-
.
.
.
.
.
0.0003
0.0003
-
.
0.020
.
0.012
-
-
.
.
.
0.020
.
. 0.015
0.044
-
-
•
Sum HyalfUa Qiirononuu
azteca tentara
0.040
0.360
0.293
0-073
0.097
0.058
0.180
0.115
-
-
0.074
-
0.025
0.173
- -
0.145
-
0.123
.
.
-
_
.
. .
.
.
0.167
0.011
.
.
0.070
.
0.088
-
.
-
-
.
0.144
.
0.033
0.075
.
.
-
92.5
90
92.5
100
0
97.5
92.5
95
95
77.5
97.5
-
97.5
96.5
90
too
100
95
97.5
97.5
75
97.5
.57.5
72.5
95
. .
35
75
80
97.5
97.5
97.5
100
95
95
95
-•
60
97.5
90
62.5
75
100
82.5
*
70
40
90
90
97.5
90
100
100
87.5
100
87.5
too
- -
97.5
92.5
87.5
100
100
100
97.5
97.5
92.5
100
65
57.5
90
.
35
72.5
82.5
100
97.5
95
100
90
100
92.5
.
55
100
95
65
95
55
72.5 •
.
67.5
aAVSLOD=0.05umS/g
b Insufficient pore-water volume for metals analysis
c Cadmium LOD=0.01 ug/L (0.0002 IWCTU)
er LOD =0.2 ug/L (0.0003 IWCTU)
LOD=0.1 ug/L (0.0001 IWCTU)
f Nickel LOD =0.5 ug/L (0.0005 IWCTU)
Source: Columns for Sample, TOC. SEM, AVS. SEM-AVS and IWCTU taken directly from Leonard et al.. !996a. Column for survival
from personal communication with Leonard. 1998.
-------
APPENDIX C
-------
APPENDIX D
Concentrations of SEM, AVS. Toxicity and TOC for EMAP. NOAA NS AT and REMAP Databases
STUDY'
._
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
EMAP-VA
SEM
_*"H2!/g.__
.289
1.500
.066
.134
.266
.266
1.292
.347
.750
.212
.497
.624
.032
.988
.604
.031
1.597
1.065
.189
.018
.079
.421
.798
.903
1.202
,159
.246
.687
.699
1.663
.083
.740
.878
.044
.910
.567
.734
2.171
3.423
.197
.162
2.803
.472
2.079
.445
2.228
.847
1.402
1.425
.263
2.936
.394
3.074
2.555
.452
.173
.578
.209
AVS
__umoj£
1.400
.742
.029
.028
3.740
1.080
1.230
.087
.948
.283
.490
13.400
.024
81.100
3.340
.331
72.400
. 8.480
6.460
.034
.976
3.210
68.000
3.150
67.700
3.310
4.870
2.420
.430
116.000.
1.300
.976
1.220
.025
3.430
.621
25.000
5.610
138.000
.892
3.590
11.900
12.500
26.600
.056
15.100
17.300
52.700
22.300
.079
29.600
.031
10.400
.402
.480
.201
.257
•3.460
SEM-AVS
ugnol/g
-I. Ill
.758
.037
.106
-3.474'
-.814
.062
.260
-.198
-.071
.007
-12.776
.008
-80.112
-2.736
-.300
-70.803
-7.415
-6.271
-.016
-.897
-2.789
-67.202
-2.247
-66.498
-3.151
-4.624
-1.733
.269
-114.337
-1.217
-.236
-.342
.019
-2.520
-.054
-24.266
-3.439
-134.577
-.695
-3.428
-9.097
-12.028
-24.521
.389
-12.872
-16.453
-51.298
-20.875
.184
-26.664
.363
-7.326
2.153
-.028
-.028
.321
-3.251
SURVIVAL"
%
100.
98.
99.
103.
99.
102.
107.
102.
99.
108.
103.
113.
101.
101.
107.
98.
102.
93.
103.
99.
97.
111.
104.
99.
105.
104.
106.
93.
91.
100.
99.
101.
98.
106.
104.
104.
107.
102.
100.
107.
82.
101.
101.
94.
106.
103.
99.
109.
88.
84.
100.
87.
104.
96.
100.
98.
101.
96.
SIGNIFICANCE1
%
0
n
V
o
V
0.
0
0
0
o
V
o
V
0
0
0'
. • o
0
0
0
0
A
V
A
\J
.
0-
0
•
0
0
0
-
0.
_
0
TOC
£n
.60
2.68
*n
»ny
ff
.56
Ion
•9U
irt
.30
A<
.95
**•
.37
1.00
1.58
„ 3.36
1<3O
.38
AQ
.09
4IO
,19
3 IT
.1 /
1*>
.32
tc
.13
1 A
.14
AQ
.4V
2.84
2 AC
.oj
2.28
£ 1
.51
*ji
./I
1it\
. l(J
2.05
41 f
.12
1A'
.14
2.30
2.84
| £
.13
3.00
.76
2*1 1
.21
2.57
4t A '
.14
n
.37
O i
.01
21*.
.36
T T?
2. / /
3| Q
. 10
1A
.20
2m
.92
2*4Q
.Jo
2.70
34 J
.14
^n
.27
4.15
1 O
.18
2.47
2.18
1.07
"I"}
.22
£.t
.65
.36
-------
APPENDIX D
-------
STUDY*
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA-HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA-HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- HR
NOAA- BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
SEM
_____umql/£
1.753
2.447
1.839
1.296
1.697
1.390
2.310
.399
2.481
1.736
.958
9.192
1.525 '
.678
5.037
4.202
1.174
1.855
3.092
2.997
2.581
2.869
5.442
2.618
5.061
2.376
6.998
4.480
4.662
5.896
3.103
1.662
3.512
.273 .
.335
1.664
2.674
5.532
4.029
4.614
3.379
4.240 '
4.303
5.209
4.801
4.697
2.600
1.013
J.527
.505
3.341
3.449
.270
.341
.888
.722
.362
2.138
3.008
.151
AVS
.• SHIP!'
17.697
10.958
68.306
56.838
9.089
43.801
51.857
3.899
19.604
148.969
18.622
120.622'
81.842
5.679
69.320
21.980
27.540
14.170
51.770
79.710
61.050
28.080
25.900
L080
12.240
4.390
63.450
20.780
23.720
51.580
59.780
7.230
25.840
.050
.036
18.760
3.630
29.210
18.440
20.530
30.120
19.320 .
22.570
14.570
35.370
54.710
56.730
10.160
15.130
.630
43.920
37.860
.950
.156
12.971
4.948
.936
3.295
3.941
.555
SEM-AVS
— _HKS!/g_
-15.944
-8.511
-66.467 '
•55.542
-7.392
•42.411
•49.547
-3.500
-17.123
-147.233
-17.664
-111.430
-80.317
-5.001
-64.283
-17.778
-26.366
-12.315
-48.678
' -76.713
-58.469
-25.211
' -20.458
1.538
-7.179
-2.014
-56.452
. -16.300
-19.058
-45.684
-56.677
-5.568
-22.328
.223
.299
-17.096
-.956
-23.678
-14.41!
-15.916
-26.741
-15.080
-18.267
-9.361
-30.569
-50.013
-54.130
-9. 147
-13.603
-.125
-40.579
-34.4H
. -.680
.185
-12.083
-4.226
-.574
-1.157
-.933
-.404
SURVIVAL"
%
94.
94.
95.
96.
97.
97.
97.
99
99.
99.
99.
100.
102.
103.
o.
41.
11.
18.
101.
112.
119.
81.
95.
109.
97.
108:
0.
20.
14.
2.
77.
19.
0.
91.
93.
69.
3.
96.
51.
91.
88.
101.
102.
101.
70.
38.
37.
29.
68.
105.
86.
76.
96.
84.
92.
85.
98.
95.
• 95. .
96.
SIGNIFICANCE8
___J6
Q
0
o
0
0
o
V
o
V
o
o
0
0
0
o
V
1
1 •
I
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1
1 •
1
4 .
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o
V
0
o
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o
\f
0'
I
. * «
1
1 .
0
\f
1
•1 .
o
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n
V
o
xf
o
u
1
J .
t
1 .
1
1 .
1
1 t
1
I .
o
V
n
u
1
0.
•
0
\J
0-
_
0
0
V
0
TOC
]A 1
.41
4.45
2.54
3nc
•U3
2.68
•» 97
j.&l
3 -lr
.JJ
art
.oil
311
.Ji
2QA
.y*
1.77
4A1 •
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2.96
1A<
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3 At
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1 99
I.oo
4AA
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38£
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3.09
2O£
.50 .
2 en
.Ju
2*jn
.zu
2.67
2 Off
.70
2AQ
.47
t Oft
4 <75
2QC
.Vo
31Q
• 47
4 tut
./O
3 no
.77
2jCt
.01
4AA
.44
.(//
.VI
AQ
.07
t nf\
1 'UU
31 S
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2-jrt
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I QA
1 .7**
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.ov
31 <
. I J
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z.yo
3 A *?
.**!
1 ,17
1 .4 /
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TC
.Z3
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1 A-l
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,zo
AA
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4 AC
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Aft
.4U .
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.zo
«"
1 9
. lo
.15
-------
STUDY1 SEM AVS SEM-AVS
IM_ HnSL^^JJSSSKBL
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA-U
NOAA- LI
NOAA- LI
NOAA-U
NOAA-U
NOAA- LI
NOAA-U
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA-U
NOAA- LI
NOAA-U
NOAA-U
NOAA- LI
NOAA-U
NOAA- U
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- U
NOAA- LI
NOAA- U
NOAA- U
NOAA- U
NOAA- U
NOAA- U
' NOAA- U
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- LI
NOAA- U
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NNOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
NOAA- BO
2.813
1.235
2.198
3.624
3.594
1.342
2.462
.964
.332
2.311
.623
.896
.544
.641
.355
.222
2.262
1.307
1.963
2,785
4.333
1.927
.004
3.831
.808
1.783
2.622
.597
1.181
1.862
2.726
2.102
2.471
1.870
1.607
4.942
2.705
2.087
1.514
2.629
3.194
.872
1.080
.123
2.914
2.218
2.609
3.650
1.634
1.267 .
2.892
2.511
.661
2.458
1.872
.959
2.480
.784
.943
1.683
35.050
2.080
14.690
21.800
27.410
37.970
46.450
1.000
4.010
79.890
6.610
16.370
2.170
2.060
1.390
4.180
39.960
.380
51.820
61.020
16.080
3.710
24.580
9.250
.960
40.630
61.840
1.090
3.730
50.390
62.760
33.630
7.220
17.120
17.810
100.800
83.010
26.730
30.880
32.050
35.390
25.810
11.300
5.310
2.893
2.369
43.959
101.984
5.237
3.256
80.584
2.241
13.490
23.077
48.062
53.288
7.599
22.486
8.831
42.399
-32.237
-.844
-12.492
-18.176
-23.816
-36.628
-43.988
-.036
-3.678
-77.579
-5.987
-15.475
-1.626
-1.419
•1.035
-3.958
-37.698
.927
-49.857
-58.235
-11.747
-1.783
-24.576
•5.419
-.152
-38.847
-59.218
-.493
-2.549
-48.528
-60.034
-31.528
•4.749
-15.250
-16.203
-95.858
-80.305
-24.643
-29.366
-29.421
-32.196
-24.938
-10.220
-5.187
-021
-.151
-41.350
-98.334
-3.603
-1.989
-77.692
.270
•12.829
-20.619
-46.190
-52.329
•5.119
•-21.702
-7.888
•40.716
SURVIVAL"
%
86.
"84.
84.
83.
82.
82.
82.
81.
81.
81.
80.
80.
79.
79.
79.
77.
77.
76.
76.
76:
75.
75.
74.
73.
71.
70.
70.
69.
68.
67.
67.
64.
63.
61.
59.
54.
53.
47.
.42.
39.
37. '
34.
16.
10.
8.
15.
26.
29.
36.
52.
83.
86.
87.
87.
89.
90.
90.
91.
91.
92.
SIGNIFICANCE1
%
0
0
0
0
0
0
0
0
0
0
0
0
TOC
0
0
0
0
0
0
0
0
0
0
3.83
1.58
2.80
2.48
2.59
1.85
3.18
1.60
1.29
3.69
.67
1.11
.27
1.56
.64
.45
2.67
1.56
3.46
3.81'
3.48
1.60
2.87
3.08
1.19
2.50
3.49
.76
.91
2.81
2.81
3.42
2.80
3.29
2.07
3.15
3.62
3.45
2.69
2.68
3.17
1.83
1.91
.22
3.05
2.89
3.74
1.83
1.72
1.53
6.98
2.12
1.00
3.15
3.25
2.39
4.45
1.88
1.78
3.41
-------
STUDY*
SEM
AVS SEM-AVS SURVIVAL' SIGNIFICANCE* TOC
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
- REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-NB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-Rfl
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
REMAP-RB
1.894
3.149
.632
1.057
.638
1.087
3.711
2.990
8.894
1.277
3.925
5.632
6.809
7.645
4.012
3.905
.942
3.515
2.216
3;323
3.391
• 3.443
2.466
2.294
5.768
1.013
2.479
.554
5.222
5.116
14.791
4.917
.398
4.855
3.290
5.822
9.167
6.214
.794
4.985
5.280
2.268
6.678
2.833
.333
.756
• .582
1.012
1.596
.326
2.709
5.485
3.596
5.329
.337
.986
.856
' 5.364
1.706
" .371
25.394
64.643
1.310
4.647
.218
.312
17.184
59.256
60.816
23.266
42.727
114.770
135.354
150.012
43.663
26.229
6.531
7.134
11.243
7.573
4.820
3.982
20.273
11.046
5.028
11.079
25.687
2.634
22.617
7.352
109.780
. .530
.218
9.606
10.105
51.460
93.563
42.415
2.651
43.663
1.934
6.300
17.559
45.222
22.315
1.216
.821
.567.
.447
.156
3.120
14.666
19.503
4.321
2.901
.156
.156
' 39.700
23.515
4.210
-23.500
-61.494
-.678
-3.590
.420
.775
-13.473
-56.266
-51.922
-21.989
-38.802
-109.138
-128.545
-142.367
•39.651
-22.324
-5.589
-3.619
-9.027
-4.250
-1.429
-.539
-17.807
-8.752
.740
-10.066
-23.208
-2.080
-17.395
-2.236
-94.989
4.387
.180
-4.751
•6.815
-45'.638
. -84.396
-36.201
-1.857
-38.678
3.346
-4.032
-10.881
-42.389
-21.982
-.460
-.239
.445
1.149
.170
-.41!
-9.181
-15.907
1.008
-2.564
830
* OJ\J
700
• ' W
-34.336
-21.809
-3.839
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92.
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91.
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96.
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96.
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VJ .
93.
91.
%
0
0
0
0
0
0
0
0
0
0
0
0
0
0
-0
0
0 .
0
0
0
0
0
0
1.
0
0
0
0
1.
1.
0
0
.
1.
1.
1.
1.
0
1.
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1.
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0
0
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0
0
.0
1.
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- 1.
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0
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' 0
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0
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.98
.90
1.51
2.44
3.52
7.36
3.99
5.24 .
3.63
3.18
3.85
4.29
4.36
6.04
3.73
3.93
.67
.75
1.22
1.25
1.05.
.88 •
1.40
.95
1.77
.76
.99
.60
1.48
1.45
9.15
3.10
2.42
2.62
5.70
2.22
6.48
3.24
2.36
3.90
6.10
1.99
15.20
2.02
1.23
.33
.30
.30
.17
.08
.42
2.29
.88
.97
.53
.12
.51
1.17
3.21
3.54
-------
STUDY"
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-BA
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-JB
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
REMAP-LS
SEM
UIDOl/g
.115
.543
.103
.167
.073
.294
.120
.109
.185
.120
.347
.120
2.275
.344
.258
.119
.258
.494
.109
.266
.327
.230
2.026
14.550
3.332
3.763
.357
.524
.244
1.247
2.478
1.744
.131
.846
4.399
3.884
.673
3.150
.270
.162
2.880
.323
.413
.377
.099
1.100
.209
.213
.954
2.759
.71]
1-915 .
2.186
2.480
.606
3.289
3.241
.616
1.506
2.485
AVS
umol/
156
« uv
.156
.156
.932
.156
.156
.156
.156
.156
.156
.156
.156
16.592
.012
.343
.156
.156
.156
.156
.156
.393
6.400
47.793
389.857
243.322
201.687
10.923
3.974
4.502
48.130
47.376
.156
1.184
.927
116.954
237.650
21.769
43.975
4.491
.873
153.755
1.684
3.056
3.056
.686
58.945
1.466
.780
•1.542
6.498
10.240
12.596
17.605
23.523
2.501
91.773
56.100
1.070
26.201
28.248
SEM-AVS
ugmol/g
~-m«w» v^w wta JMhm,m • *•
f\A t
-.041
.387
-.053
-.765
-.083
.138
-.036
-.047
.029
-.036
-.036
-14.317
.332
-.085
-.037
IfYt
-1U£
.338
-.047
i in
.110
-.066
-6.170
-45.767
-375.307
-239.990
-197.924
-10.566
-3.450
-4.258
-46.883
-44.898
1.588
-1.053
-.081
-112.555
-233.766
-21.096
-40.825
: -4.221
•7 t |
-.'11
-150.875
-1.361
-2.643
-2.679
^7
* J .O /
f*-j
-. JO/
COD
*.JoH
-3.739
-9.529
-10.681
-15.419
-21.043
-1.895
-88.484
-52.859
jtejt
-.434
-24.695
-25.763
SURVIVAL"
%_
99.
94.
85.
97.
99.
91.
84.
92.
90.
88.
89.
81.
69.
91.
94.
84.
91.
86.
89.
86.
93.
83.
51.
0.
37.
79.
95.
98.
84.
91.
36.
69.
94.
73.
93.
• 89.
77.
91.
91.
98.
92.
93.
94. ,
92.
93.
96.
93.
95.
83.
96.
97.
97.
'95.
99.
98.
95.
97.
95.
96.
96.
SIGNIFICANCE*
%
0
0
0
0
0
0
0
0
0
0
0
0
1.
0
0
0
0
0
0
0
0
0
1.
1.
1.
1.
0
0
0
0
1.
1.
0
1.
0
0
1.
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
TOC
.08
.07
.05
.16
.05
.34
.83
.92
4.48
.83
1.26
.62
1.81
3.85
.77
2.23
.88
2.10
4.07
1.06
.29
.19
.77
1.52
.83
.97
.26
.35
.27
.54
1.12 '
1.14
.21
1.58
6.55
8.45
4.11
5.47
.74
1.40
7.70.
.20
1.20
1.30
.75
3.86
.58
.69
.26
.45
.56
.21
- .27 ,
.32
.25
.77
1.14
.15
.95
.25
•
.
-------
-------
STUDY'
SEM
SURVIVAL*
SIGNIFICANCE" TOC
REMAP-RB .193 .156 .037
REMAP-RB .869 19.617 -18.748
REMAP-RB 1.288 .593 .695
REMAP-RB 1.650 .624 1.026
REMAP-RB 2.422 .156 2.266 •
REMAP-RB .512 .156 .356
REMAP-RB 4.198 4.086 .112
REMAP-RB 5.081 36.490 -31.409
REMAP-RB 6.095 . 5.957 .138
REMAP-RB 8.471 8.078 .393
REMAP-RB 3.370 17.247 -13.877
REMAP-RB 1.198 .156 1.042
REMAP-UH 2.127 12.446 -10.319
REMAP-UH 1.360 1.790 -.430
REMAP-UH 1.197 3.373 -2.176
REMAP-UH 1.975 17.136 -15.161
REMAP-UH 2.829 25.189 -22.360
REMAP-UH 2.830 56.401 -53.571
REMAP-UH 1.385 44.588 -43.203
REMAP-UH 1.519 11.549 -10.030
REMAP-UH 3.186 86.235 -83.049
REMAP-UH 2.086 11.713 -9.627
REMAP-UH 1.799 12.631 -10.832
REMAP-UH .930 10.093 -9.163
REMAP-UH .459 .156 .303
REMAP-UH .889 2.623 -1.734
REMAP-UH .833 2.464 -1.631
REMAP-UH 1.317 15.563 -14.246
REMAP-UH 2.480 32.123 -29.643
REMAP-UH .626 9.949 -9.323
REMAP-UH 1.500 5.427 -3.927
REMAP-UH .723 1.341 -.618
REMAP-UH 4.158 13.504 -9.346
REMAP-UH 2.241 27.788 -25.547
REMAP-UH 2.907 29.285 -26.378
REMAP-UH .852 1.591 -.739
REMAP-UH 2.294 53.955 -51.661
REMAP-UH 2.995 33.995 -31.000
REMAP-UH 2.981 44.910 -41.929
REMAP-UH .677 10.323 -9.646
a) Sources: EMAP-VA is U.S. EPA. 1996
NOAA-L1 is Wolfe et at., 1994
NOAA-BO is Long et al., 1996
NOAA-HR is Long et al.. 1995
REMAP is Adams et al.. 1996
b) Conclusion of signtfigance varies for three databases.
EMAP significance based on percent survival of control
NOAA significance based on percent survival less than 80%
REMAP significance based on percent survival less than 80%
c) Significance: 0 - No significant toxicity
1 - Significant toxicity
rv
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