NOAA Technical Memorandum NOS ORCA 87

 

The Utility of AVS/EqP in Hazardous Waste Site Evaluations

Proceedings of a workshop sponsored by the Coastal Resource Coordination Branch,

Hazardous Materials Response and Assessment Division, NOAA

Seattle, Washington

August 1995

noaa NATIONAL OCEANIC AND ATMOSPHERIC ADMINISTRATION

____________________________________________________________________________________________

National Ocean Service
Office of Ocean Resources Conservation and Assessment

National Ocean Service

National Oceanic and Atmospheric Administration

U.S. Department of Commerce

The Office of Ocean Resources Conservation and Assessment (ORCA) provides decisionmakers

comprehensive, scientific information on characteristics of the oceans, coastal areas, and

estuaries of the United States of America. The information ranges from strategic, national

assessments of coastal and estuarine environmental quality to real-time information for

navigation or hazardous materials spill response. Through its National Status and Trends

(NS&T) Program, ORCA uses uniform techniques to monitor toxic chemical contamination of

bottom-feeding fish, mussels and oysters, and sediments at about 300 locations throughout the

United States. A related NS&T Program of directed research examines the relationships between

contaminant exposure and indicators of biological responses in fish and shellfish.

ORCA provides critical scientific support to the U.S. Coast Guard during spills of oil or

hazardous materials into marine or estuarine environments. This support includes spill trajectory

predictions, chemical hazard analyses, and assessments of the sensitivity of marine and estuarine

environments to spills. The program provides similar support to the U.S. Environmental

Protection Agency's Superfund Program during emergency responses at, and for the cleanup of,

abandoned hazardous waste sites in coastal areas. To fulfill the responsibilities of the Secretary

of Commerce as a trustee for living marine resources, ORCA conducts comprehensive

assessments of damages to coastal and marine resources from discharges of oil and hazardous

materials.

ORCA collects, synthesizes, and distributes information on the use of the coastal and oceanic

resources of the United States to identify compatibilities and conflicts and to determine research

needs and priorities. It conducts comprehensive, strategic assessments of multiple resource uses

in coastal, estuarine, and oceanic areas for decisionmaking by NOAA, other Federal agencies,

state agencies, Congress, industry, and public interest groups. It publishes a series of thematic

data atlases on major regions of the U.S. Exclusive Economic Zone and on selected

characteristics of major U.S. estuaries.

ORCA implements NOAA responsibilities under Title II of the Marine Protection, Research, and

Sanctuaries Act of 1972; Section 6 of the National Ocean Pollution Planning Act of 1978; the Oil

Pollution Act of 1990; the National Coastal Monitoring Act of 1992; and other Federal laws. It

has four major line organizations: Coastal Monitoring and Bioeffects Assessment Division,

Hazardous Materials Response and Assessment Division, Strategic Environmental Assessment

Division, and the Damage Assessment Center.
NOAA Technical Memorandum NOS ORCA 87

The Utility of AVS/EqP in Hazardous Waste Site Evaluations

September 13-15, 1994

Woods Hole, Massachusetts

Edited By:

Donald A. MacDonald

NOAA Hazardous Materials Response and Assessment Division

7600 Sand Point Way Northeast

Seattle, Washington 98115

Sandra M. Salazar

E.V.S. Consultants, Inc.

200 West Mercer Street

Seattle, Washington 98119

Seattle, Washington

_________________________________________________________________________________________________________________________________________________________________________________________________________________________________________

United States National Oceanic and National Ocean Service

Department of Commerce Atmospheric Administration W. Stanley Wilson

Robert H. Brown D. James Baker Assistant Administrator

Secretary Under Secretary for Oceans for Ocean Services and

and Atmosphere Coastal Zone Management

_________________________________________________________________________________________________________________________________________________________________________________________________________________________________________
Hazardous Materials Response and Assessment Division

Office of Ocean Resources Conservation and Assessment

National Ocean Service

National Oceanic and Atmospheric Administration

U.S. Department of Commerce

Silver Spring, Maryland

NOTICE

This report has been reviewed by the National Ocean Service of the National Oceanic and Atmospheric

Administration (NOAA) and approved for publication. Such approval does not signify that the contents of

this report necessarily represent the official position of NOAA or of the Government of the United States,

nor does mention of trade names or commercial products constitute endorsement or recommendation for

their use.
Contents

Introduction............................................................................................................................1

NOAA’s Role in the Superfund Process...........................................................................3

Alyce Fritz

Opening Comments..............................................................................................................5

Donald MacDonald

Assessing the Assessment Tools........................................................................................7

Robert Dexter

An Introduction to Equilibrium Partitioning Applied to

Nonionic Organic Chemicals and Metals.......................................................................11

Dominic DiToro

Generic Screening versus Site-Specific Reality: What Works,

What Doesn’t, and Why?...................................................................................................19

Peter Chapman

Certain Chemical Aspects of Heavy Metal-Sediment

Interactions: The Role of Organic Carbon and the Kinetics

of Metal Sulfide Oxidation.................................................................................................23

John Mahony

The Influence of Sediment Stability and Disturbance on

Marine Amphipod Toxicity...............................................................................................27

Jack Word

KOC and KDOC in Sediment Pore Water........................................................................33

Jim Brannon

Pore Water Toxicity Testing: A Direct Measure of EqP...............................................37

Robert Scott Carr

Predictions of Cadmium Accumulation in Benthic Organisms...............................41

André Tessier

Predicting the Toxicity of Metals-contaminated Sediments using

AVS and Interstitial Water Normalizations: Results from Spiked

Sediment Laboratory 10-day Lethality and 4-month Colonization

Tests.......................................................................................................................................45

David Hansen

Marathon Battery: What Were the Questions?...........................................................51

Mark Sprenger
Contents continued

Use of SEM/AVS Ratios and Pore Water Metal Concentrations

to Predict Metal Bioavailability in Field Sediments.....................................................55

Gary Ankley

Sediment Contamination, Toxicity, Toxicity and Macrobenthos:

Comparability of Laboratory and Field Observations..................................................61

Rick Swartz

Comparison Between the Results of Toxicity Testing with Marine Sediments and

Predicted Effects Based on Chemistry..............................................................................67

Jack Anderson

Assessing Toxicity in Multi-Metal Contaminated Sediment:

Application of Theoretical and Investigative Approaches.........................................69

Michael Johns

Strengths and Weaknesses of the EqP Approach as Applied

to Hazardous Waste Sites Risk Assessment..................................................................73

Spyros Pavlou

Panel I Discussions..............................................................................................................77

Panel II Discussions.............................................................................................................81

Panels III and IV Discussions............................................................................................87

Panel V Discussions............................................................................................................93

NOAA’s Overview and Summation.............................................................................97

Appendix A Workshop Participants...........................................................................A-1

Appendix B Agenda.........................................................................................................B-1

Appendix C References....................................................................................................C-1

Appendix D Acronyms and Abbreviations................................................................D-1

ii
Introduction

To aid in their understanding of contaminant behavior at Superfund sites, the Coastal

Resource Coordination Branch (CRCB) of the National Oceanic and Atmospheric

Administration (NOAA) organized and chaired a workshop entitled “The Utility of

AVS/EqP in Hazardous Waste Site Evaluations.” Twenty-four invited scientists, active in

the field of sediment evaluation, along with the CRCB staff and representatives of the U.S.

Environmental Protection Agency (EPA) and the U.S. Fish and Wildlife Service, attended

the workshop held September 13-15, 1994, at the Marine Biological Laboratory in Woods

Hole, Massachusetts. A complete listing of participants is provided in Appendix A.

The purpose of the workshop was threefold:

(1) Clarify the scientific theory on which the equilibrium partitioning (EqP) and acid

volatile sulfides (AVS) methodologies for sediment evaluation are based.

(2) Determine what the EqP and AVS methodologies can and cannot tell us about

sediment contamination and toxicity.

(3) Identify how these methodologies can be applied to the evaluation of hazardous

waste sites.

Individual presentations and panel discussions were used as vehicles of information

transfer. Overall, 15 presentations were given by leaders in the fields of EqP and AVS, and

four panel discussions were convened with each panel addressing three specific questions,

except for Panel III that addressed six questions (Panel III was formed by merging two of the

original panels). Panels consisted of three to five participants, with each participant

bringing a specific element of expertise to the panel. While a formal agenda (Appendix B)

had been prepared for the workshop, to facilitate and maximize the free exchange of ideas,

discussions generated by either the individual presentations or the panels were generally

allowed to run their course. The unstructured atmosphere promoted uninhibited discussions

and helped attain consensus on several issues.

The following proceedings are an attempt to capture the tone and level of enthusiasm

generated by workshop participants. Summaries are provided of each presentation and

panel discussion. In preparing these summaries, an attempt was made to not only

summarize the facts presented by the speaker(s) but also, to provide a sense of the overall

workshop atmosphere. Audience members were encouraged to express their interpretation

of the data being presented and ask questions throughout the sessions. Therefore, when

appropriate, the presentation summaries include summaries of these ensuing discussions,
AVS Proceedings

indicated by [ ] and italics. Brackets and italics are also used for editorial clarifications for

the reader. When multiple authors are listed for a presentation an asterisk (*) is used to

indicate the presenting author.

Since this document constitutes only a summary of the workshop presentations and does

not include any of the data or graphics included in the presentations, it is strongly

recommended that if the reader has any questions or if anything is unclear, he should refer

to the literature, contact the presenter directly, or contact Don MacDonald, Workshop

Coordinator and Chairman (E-mail: Don_MacDonald@hazmat.noaa.gov) for more

information. To help the reader, a list of references, by author, is included as Appendix C.

This list includes articles by the workshop participants, some of which contain the data on

which the presentations were based; articles cited in the presentations; as well as, other

articles dealing with EqP and AVS. This reference list is not meant to be all inclusive on

these topics but a starting point for anyone interested in more detailed information on these

subjects.

NOAA would like to extend a sincere thank you to all attendees for their participation in

this workshop. The information generated though the presentations, panel discussions, and

question and answer sessions is extremely valuable to the CRCB in our evaluation of

Superfund sites. We encourage additional discussions and research in this area and would

be willing to serve as a repository of data collected in an effort to better understand the

chemical and biological processes that regulate toxicity at aquatic sites.

2
AVS Proceedings

NOAA’s Role in the Superfund Process

Alyce Fritz

CRCB, HAZMAT, NOAA

Seattle, Washington

NOAA serves as a Federal trustee in the protection and management of natural resources.

Trusteeship extends to natural resources that are found in or under waters navigable by

deep draft vessels, in or under tidally influenced waters, waters of the contiguous zone, the

exclusive economic zone, and the outer continental shelf, and in inland areas serving as

habitat for marine mammals and other protected species. Examples of resources managed

under trusteeship include marine fishery resources and their supporting ecosystems,

anadromous fish, endangered species, and marine mammals. The Hazardous Materials

Response and Assessment Division (HAZMAT) is charged with applying science and

technology to problems resulting from releases of oil and hazardous substances to protect

the public interest. The primary goal of HAZMAT is to work hand-in-hand with other

trustees to reduce risks to coastal habitats and resources through preparedness and

response activities. Technical support is provided to EPA regional offices for evaluating

natural resource concerns at hazardous waste sites. The program works to improve

coordination with Federal and state natural resource trustee agencies under its

Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) (i.e.,

Superfund) responsibilities.

NOAA’s interests are pragmatic—NOAA must address resources and contaminant issues

for nearly 500 sites, 30 of which are high priority sites. To provide technical support for

environmental investigations to determine ecological risks, NOAA must continuously draw

from the “environmental toolbox”. While many of the new research results are based on

sound science, their utility in correctly assessing hazardous waste sites is often unclear. The

utility of AVS and EqP methods in hazardous waste site evaluations is of primary interest

to NOAA because it is necessary to not only know what contaminants are present, but also

what portion is bioavailable. NOAA has specific questions regarding the application of the

AVS/EqP approach: How should it be used with mixtures of contaminants? In what

phase of the evaluation process should AVS and/or EqP information be used? What is the

recommended approach for ecological risk assessment at hazardous waste sites? NOAA’s

primary goal for this workshop is to reach consensus on technical issues associated with the

AVS and EqP approaches and identify ways that this type of data can best be applied to

risk assessment.

3
AVS Proceedings

4
AVS Proceedings

Opening Comments

Donald MacDonald

CRCB, HAZMAT, NOAA

(DON_MACDONALD@HAZMAT.NOAA.GOV)

Seattle, Washington

The purpose of the AVS/EqP Workshop is to review the tools available to NOAA to fulfill

their mandate. Specifically, how can the AVS and EqP approaches be used by NOAA as

they recommend and provide comment on sampling programs to evaluate conditions at

hazardous waste sites. What type of information can the various methodologies provide?

What are the most appropriate tools to use to evaluate hazardous waste sites? If the AVS

and EqP approaches are appropriate tools, what caveats apply to their use? Many times

the remedial activities are more destructive to habitats and trust resources than if the

contaminants were left in place. To ensure optimum management of contaminated

environments, NOAA must apply the appropriate tools during data evaluation. The task at

hand is not an easy one, particularly when multiple contaminants are at issue.

5
AVS Proceedings

6
AVS Proceedings

Assessing the Assessment Tools

Robert Dexter

E.V.S. Environmental Consultants

Seattle, Washington

Over the past 20 years, tremendous resources have been dedicated to identifying problems

associated with toxic substance releases. In these efforts, cost per value has been a major

consideration. In other words, studies that do not support management decision-making

are not looked on favorably, regardless of cost. In contrast, since even the relatively most

expensive assessment costs are only a small proportion of possible (or actual) clean-up

costs, if the work will assist in the decision-making process, it will usually be approved even

if it seems very expensive in comparison to what the scientific investigators are used to.

Many of the assessment tools tend to be precise or accurate, but not both. For example,

chemical analyses are quite precise, but the analytical results are not good estimates of

contaminant bioavailability. On the other hand, biological measurements and bioassays are

accurate predictors or assessors of effects to biological communities but because of the

normal high level of variation in natural organisms and the difficulty of transferring

laboratory information to the field, they are not precise.

The two techniques to be discussed over the next few days are intended to address some of

those limitations. These procedures are of interest because they offer hope that the accuracy

of assessments of the impacts of releases of toxic substances associated with sediments can

be improved without large increases in the basic costs of the assessment and without the

need to create and prove the utility of a wholly new technological approach.

Three major factors should be considered in this workshop on the utility of AVS and EqP

approaches: good science, applicability to the real world, and the potential to support

management decisions. AVS and EqP are based on fundamental chemical principals and

good science; partitioning does occur and sulfides do affect the bioavailability of some

metal contaminants. These principals support the contention that the availability of

substances in the sediments is primarily a function of only a few, easily measured,

parameters. However, it may not be possible to accurately translate these principals to

situations in the field. Both approaches are based on thermodynamics and are expected to

provide accurate estimates when steady-state conditions exist, but, in the real world,

steady-state conditions are not always applicable. Secondly, the bioavailability of

contaminants is dependent on characteristics that vary from time to time, and while we

7
AVS Proceedings

know what most of those characteristics are it is not clear that we have the capability to

accurately measure all the factors influencing this bioavailability. For example, to what

extent is a simple measurement of bulk total organic carbon (TOC) representative of the

absorptive capacity of the sediments. The following key question should be asked when

evaluating these assessment tools: When does it become more difficult and expensive to

collect AVS and EqP data than to use alternative, direct measurement methods (e.g.,

toxicity tests)? The answer to the question of whether these approaches can be used in the

real world is the primary focus of this workshop.

While focusing on the real-world applicability of these approaches, we can’t lose sight of

how they can be used in the context of a risk-management decision. Two difficulties exist

with respect to AVS and EqP data supporting management decisions:

(1) To what extent are there technical difficulties with the application of AVS and

EqP data?

(2) What is the mechanism for incorporating such information into the decision making

process?

At the present time, the process for incorporating risk assessment, especially ecological risk

assessment, is not well established, and this lack of an a priori method to incorporate this

type of data frequently initiates additional debate rather than aiding in decision making.

As an example, while developing a work plan for an ecological assessment at a site, the

potential responsible party’s (PRP) lead ecological scientist agreed to include toxicity testing

but was positively opposed to including benthic community studies. The reason for his

attitude being that at another site benthic community studies had been performed and the

data demonstrated significant differences between the reference station and the site, but

because no determination had been made in advance as to what type of effects would be

considered important enough to warrant remediation, it just resulted in endless debate on

the meaning of the data.

The ultimate bottom line is: In what way can AVS and EqP data be used to further the

identification of seriously contaminated sediments? In this regard we have to keep in mind

that our remediation options are presently very limited: the contaminated sediments need to

be isolated from biota and that basically means either removal (i.e., dredging) or capping.

Our criteria for taking these actions is essentially the toxicity of the sediments to biota. We

currently have some direct measures of this toxicity (e.g., toxicity tests) as well as some

direct chemical measurement guidelines (e.g., Puget Sound apparent effects thresholds

[AETs], effects range low [ERLs], effects range medium [ERMs]). The question is: how will

8
AVS Proceedings

EqP and/or AVS approaches improve our capability for determining sediment toxicity?

Specifically, the following questions are still unanswered and should form the basis for

assessing the usefulness of AVS and EqP:

Can AVS and EqP data:

• replace other more expensive or complex methods of measurement?

• estimate non-measurable effects, such as those to higher trophic level organisms

caused by biomagnification?

• identify “problem” substance(s) under mixed waste exposure conditions so

cleanup can be focused on the problem substance(s)?

• explain/support other biological and chemical measures?

In summary, the two general questions which should be addressed by this workshop are:

Do these techniques tell us what they claim? and Where do these approaches fit into the

EPA decision-making process?

[Comments from the audience dealt with the importance of remembering the

biota, since they are what we are concerned with. When asked “If we are

concerned with bioavailability why not measure contaminants in the biota?” Dr.

Dexter responded that organisms suitable for bioaccumulation studies are not

always available at the study site.]

9
AVS Proceedings

10
AVS Proceedings

An Introduction to Equilibrium Partitioning

Applied to Nonionic Organic Chemicals

 and Metals

Dominic DiToro

HydroQual, Inc.

Mahwah, New Jersey

One of the major principles associated with sediment quality criteria is that bulk chemical

concentrations are useless for criteria development and clean-up levels; they are useless

because they ignore bioavailability. Bioavailability is properly accounted for by accounting

for the phase. Sediment contaminant concentrations are often reported in units of dry

weight, whereas tissue contaminant concentrations are generally reported in wet weight.

Why? By reporting sediment as dry weight, there is the implicit assumption that water is

irrelevant in sediment; that water is irrelevant to chemical activity in sediment. Why not

report sediment chemical concentrations as micrograms per liter of sediment plus water?

The correlation should be made with respect to chemical activity, as is done in chemical

thermodynamics—that is μgrams (μmoles) of chemical divided by the weight of the phase it

is in: in water, μg chemical/liter water, in air μg chemical/liter atmosphere, in sediment μg

chemical in the phase in which it occurs. Bioavailability is properly expressed by the

concentration of the chemical in terms of the phase in which it occurs, so we can get what is

called the chemical activity correct.

If dry weight is off by only a factor of two or so, which is within normal measurement error,

then dry weight would be adequate and there would be no need for the EqP and AVS

approaches. However, dry weight concentrations associated with toxicity (e.g. mortality)

can vary over orders of magnitude. We have to know where the chemicals are residing. For

nonionic hydrophobic organic compounds, organic carbon is the primary phase in

sediments. Nonionic hydrophobic organics concentrations should be normalized to organic

carbon. For metals, AVS and organic carbon are the major controlling parameters.

Biological tests are not necessarily the answer; they are not the be all and end all; they only

measure the degree of toxicity, they cannot identify the toxic component. The name of the

game we have chosen to play is “causation”. What is causing the observed bioeffects? If

you don’t know the cause of the toxicity, then you may not know the source of the toxicity.

Conducting a clean-up effort without knowing the source is dangerous because after the

11
AVS Proceedings

cleanup is complete the area may become recontaminated because you have failed to

identify and control the source of the contamination.

I want to take you through the history of the development of EqP to provide clarification of

our assumptions. It’s important to understand how and why we came to where we are.

Studies done by Bill Adams (Adams et al. 1985) on the acute toxicity of kepone to

choronimid larvae showed more than an order of magnitude difference between various dry

weight sediment concentrations associated with the toxicity. When Adams plotted the

toxicity against the corresponding pore water concentrations there was very close

agreement. Based on this analysis one would assume that pore water was the sole route of

exposure. Suppose he had plotted the organic carbon normalized sediment concentrations

against toxicity instead of the pore water concentrations; he would have found fairly close

agreement between level of toxicity and these carbon normalized concentrations. He might

then have concluded that organic carbon was the route of exposure for kepone.

Both conclusions are incorrect. You cannot tell from an equilibrated experiment the source

of the dose. EqP makes no assumptions as to the route of exposure (pore water versus

organic carbon phase); the major assumption is simply that equilibrium exists. The

equilibrium refers not to an equilibrium between the organism and the pore water or organic

carbon, but an equilibrium between the organic carbon and the pore water, such that it

doesn’t matter how the organism actually receives its dose. That is the fundamental

assumption on which EqP is built. Another basic assumption is that when the pore water

concentration is the same as the water-only concentration, the organism is seeing the same

chemical activity in the sediment/pore water exposure as it is in the water-only exposure.

This assumption says nothing about which phase is exerting the activity.

In order to test the assumption that the water-only LC50 was equivalent to the pore-water

LC50, the LC50 in water only was determined for several organic chemicals (i.e.,

phenanthrene, fluoranthene, acenaphthene, kepone, dieldrin, endrin, DDT) and tests were

run on three sediment samples by spiking them with the various compounds. The toxicity of

the various samples was determined and the corresponding pore water concentrations

measured. Then by using the concept of toxic units, it could be determined if toxicity was

properly predicted. [Pore-water toxic units (PTU) = pore-water concentration/pore-water LC50.

Predicted PTU (PPTU) = pore-water concentration/water-only LC50].

If 50 percent mortality occurs when the PPTU = 1, our assumption is correct. When the

PPTUs for all the data points were plotted against percent mortality, a reasonably good

12
AVS Proceedings

dose response curve was obtained with the 50 percent mortality point falling between about

0.6 and 3 PPTU; except for DDT, the most hydrophobic of the compounds tested. This

range is for all the data points for all chemicals. If the plot was based on calculated LC50s

rather than the individual data points, the range would be smaller. The tests supported the

assumption that you can predict the toxicity of a single contaminant in sediment based on

what you know about its toxicity in water only. We have to show that the theory works for

single contaminants before we try it with mixtures of contaminants.

If pore-water concentrations were easy to determine there would be no point in proceeding;

just use pore-water concentrations as criteria for sediments. Since that is not the case, we

need to be able to estimate pore water concentrations based on solid phase chemistry, or in

the case of establishing criteria, to estimate the solid-phase concentration that is liable to

have an adverse biological effect from the pore-water concentration known to be toxic.

A relationship between solid phase and pore water phase concentrations has been

determined for nonionic hydrophobic organic compounds: Cs=focKocCd. The organic

carbon concentration (foc) is the only sediment specific property of concern, and the organic

carbon partitioning coefficient (Koc) is the only chemical specific property of concern. So the

organic carbon normalized sediment concentration (solid phase sediment concentration (Cs)

divided by (foc) is proportional to the pore-water concentration (Cd), with the

proportionality constant being the Koc: Cs/foc=KocCd.

Therefore to establish an LC50 concentration for sediment (solid phase), you insert the

water only LC50 concentration, which is assumed to be the same as the pore-water LC50,

and multiply by the Koc. This gives you the carbon normalized LC50 for the sediment.

[A discussion ensued at this point on the variability of type of organic

carbon in sediments and its effect on the Koc value and thus on the overall

calculation. The result appeared to be that in most sediments the organic

carbon type variability is within experimental error and would probably

have little effect on contaminant availability and Koc . However, there are

exceptions, such as sediments containing; coal, wood chips, or high levels

of PAH contamination. These exceptions must be kept in mind when

analyzing any sediment data. Also, there wasn’t agreement on what

constitutes experimental error. Dr. DiToro pointed out during the

discussion that it was important to use the Koc based on the Kow

determined by the EPA Athens Lab because literature values can vary by

two orders of magnitude.]

13
AVS Proceedings

Okay, so here’s what we did to test the theory. We made a prediction of the organic carbon

normalized LC50 from the water-only LC50 by multiplying it by Koc. We then plotted our

mortality data for the seven chemicals against the predicted LC50 data. The plot was not

quite as tight as the pore-water plot. Our uncertainty analysis tells us that based on this

experimental suite of data we can predict the sediment LC50 within a factor of two. Now, I

want you to understand that this one plot tests essentially every assertion that we’ve made.

Of course, it’s only been used for lab exposures with spiked sediments, but it does test the

assertion that the water-only LC50 can be used to predict the toxicity of sediment

exposures.

[A discussion ensued around the wide range of scatter in the plot and the

fact that within the factor of two range, the mortality ranged from zero to

100 percent It was pointed out that the large scatter was partially due to the

plotting of individual test results. While within the factor of two range

for the predicted sediment LC50, there was essentially no predictability

below the range the vast majority of the results were less than 50 percent

mortality and above the range close to 100 percent mortality was observed.]

From a water-only test and a chemical property (Koc), we were able to calculate a sediment

LC50 value within a factor of two. I think that’s a very powerful demonstration of the

underlying theoretical soundness, at least within this range of experimental information. If

you have to get it greater than a factor of two, you have a real problem. I don’t think any

assessment methods are good to a factor of two.

[At this point the workshop broke for lunch. After lunch Dr. DiToro

continued his presentation addressing metals contamination and AVS.]

Early on we performed an experiment where we measured interstitial water cadmium (Cd)

concentrations with electrodes in sediments that had been spiked with various

concentrations of Cd. Plotting the interstitial water Cd concentration against sediment Cd

on a dry-weight basis shows that we only get non-detects until we hit a very high Cd

concentration. Then suddenly the interstitial water-Cd concentration jumps by two orders

of magnitude: Cd sulfide was forming while Cd was being added to the sediments. It is not

a big surprise that metals and sulfide interact in sediment; but what was remarkable was

that it was 20 or 30 μmoles per gram, which is much higher than pore-water sulfide could

possibly be.

Where was the sulfide coming from? We looked in the pore water and there wasn’t a tenth

to a hundredth of the sulfide that we needed to explain all the Cd sulfide forming. We

finally came up with the following suggestion. Dissolved Cd is in equilibrium with solid Cd

14
AVS Proceedings

sulfide (Cd+2 + S-2 CdS(s)). Iron sulfide, which is the major form of sulfide in sediments,

or at least a portion of it, namely the AVS, is also in equilibrium (FeS(s) Fe+2 + S-2).

When you add Cd to a system that is already in equilibrium, Cd and the sulfur combine,

forming Cd sulfide (Cd+2 + FeS(s) CdS(s) + Fe+2). So effectively, what we think

happens in sediments is that when heavy metals are added to the sediment they basically

titrate the AVS that is in the sediment in the form of iron sulfide. They form an insoluble

metal sulfide, which is less soluble than the iron sulfide and not bioavailable. So the

hypothesis is that there is a reservoir of sulfides in sediments. When the experiment to

measure the release of iron was done, we found a stoichiometric iron release.

Chemistry modeling indicates that regardless of pH, iron concentration, ionic strength, or

other conditions, the following is always true: The activity of the metal, the divalent activity

in the pore water over the total metal added, is always less than the ratio of the metal

solubility products ({Me+2}/[Me]A < KMeS/KFeS).

Repeating the experiment, this time keeping track of the AVS (the term AVS comes from the

geochemists, who have been using it for years; it is an operational speciation of sulfide in

sediments), we measured the dissolved metal concentration and we had nothing, nothing,

nothing and then all of a sudden metal concentration jumps up. The jump occurred when

the added metal exceeded the concentration of AVS associated with that sediment. In quite

a number of cases, the concentration jumps up later than the AVS suggesting that there are

some other binding phases in there as well. We think the next most important binding phase

is organic carbon.

So the question was, did this really work in natural sediments? Here is the operational way

to do this problem. You give us the sediment; we measure the following things: the total

amount of AVS and the amount of Cd, copper (Cu), nickel (Ni), zinc (Zn), and lead (Pb).

All of these metals are measured at the same time if they are in substantial quantities; this is

what we call the “simultaneously extracted metal”.

Now then, if the amount of metal that is extracted is less than the iron sulfide present, it

means that in that sediment there is, besides the amorphous metal sulfides, an additional

amount of AVS. Therefore, any metal that was released, in anyway, would immediately

titrate the AVS and turn back into an amorphous metal sulfide. The chemistry tells us that

an excess of AVS over the total extractable metals that can form insoluble sulfides should

result in the activity of that whole laundry list of metals being very small. The other thing

that you should realize is that you can’t treat all these metals independently. The five

15
AVS Proceedings

metals Cd, Cu, Ni, Pb, and Zn have to be treated together. There is not a criterion or an

action level for Cu. There is only an action level for the total molar concentration of all the

metals that form sulfides, more insoluble than iron sulfide. That is the fundamental reason

that I think the dry-weight metals’ criteria can’t possibly be correct in metal situations where

there is more than one metal. They are not independently applicable.

If, on the other hand, you end up with a simultaneously extracted metal concentration

greater than the AVS concentration, then you know that there is no iron sulfide in the

system. Furthermore, you know that there must be metals in the other phases and they can

then cause pore-water concentrations to increase and possibly cause toxicity. What counts

is how strong and how much of the other binding phases are present.

So, if this sulfide mechanism is correct, it says

 (1) You cannot do metals separately, you have to do them all at once.

(2) It is the AVS that controls the lack of toxicity.

Looking at a real data set, when the response data for the 17 sediments from Foundry Cove

were plotted against dry-weight concentration, there was no relationship between organism

response and dry-weight Cd and Ni. On the other hand, when the same data set is plotted

against the molar ratio of simultaneously extracted metals to AVS, you get a relationship

that tells you that if the ratio is <1 you see no mortality. However, if the ratio is >1, we

don’t really know if it will be toxic.

Question Is it more available via ingestion?

DiToro It appears not to be. I mean there are oligochaetes in that data set. The strongest

bond that a metal ion can make is to a sulfur group. There is nothing that is

chemically stronger than that. You use acid to dissolve it, basically. Quite a lot

of it won’t dissolve. It turns out that if you do a weak hydrochloric extraction,

for example Cu sulfide hardly reacts at all. It’s even stronger than the chloride

extraction. In which case the organisms haven’t a chance at getting a hold of it.

They just don’t have ligands that can complex that stuff more strongly than the

sulfides. That’s what the experiment suggests.

Question What happens in a periodically saturated environment, such as wetlands?

DiToro Wetlands. We have thought about this a lot and have misled ourselves and a lot

of people on what went on. We used to think that because iron sulfide oxidized

rapidly that was important. What really matters is how fast the metal sulfides

themselves oxidize. As far as we know right now very few of them oxidize very

quickly. My guess would be that they are very stable, with the exception of Cd

sulfide that oxidizes rapidly.

16
AVS Proceedings

17
AVS Proceedings

Generic Screening versus Site-Specific Reality:

What Works, What Doesn’t, and Why?

Peter Chapman

E.V.S. Environmental Consultants

North Vancouver, British Columbia

Canada

Sediments are a complex issue, with many processes (i.e., micro environments, redox

gradients, AVS, physical/chemical interactions, and biology) affecting contaminant

bioavailability. Biologists, who are primarily concerned with effects to the organisms, look

to chemistry for explanations of observed results, and assistance in assessing what should

be remediated. The primary question is “What are contaminated sediments?” Today,

contaminated sediments can be defined numerically or by biological effects. Site-specific

criteria are also used to define contaminated sediments. However, each approach has

management implications. The basic message of my presentation is that AVS and EqP are

“tools in the environmental assessment toolbox” that serve three useful purposes: (1)

hypothesis testing, (2) prioritization, and (3) helping explain observed effects (or lack

thereof). Case studies will be presented to demonstrate where the approach works, and

where it doesn’t.

AVS, with simultaneously extracted metals (SEM), provides a basis for predicting the lack

of bioavailability for some inorganic contaminants (i.e., when the ratio of SEM/AVS is <1).

AVS is not universally applicable; it does not work for all inorganic contaminants, and lack

of bioavailability can also occur when the ratio is >1. Existing criteria for some organic

compounds are based on EqP values. The primary concern with this approach is that EqPderived values are based on single contaminants and we are often faced with a complex

mixture of contaminants. There is no guarantee that other things in the sediment are not

also causing effects. AVS and EqP should not be used alone, as inflexible guidance (i.e., as

though set in stone) without consideration of site-specific conditions. Thought must be

applied to the evaluation of contaminated sediments.

AVS and EqP should be used primarily during level 1 of a risk assessment—as a screening

and prioritization tool to focus expertise and resources. A tiered approach is recommended

with the primary goals to (1) identify polluted sediments, (2) prioritize problems, and (3)

remediate effectively. Results from level 2 measurements should be analyzed in conjunction

with data from level 1. For example, AVS and EqP data can assist in explaining observed

18
AVS Proceedings

bioassay results and/or field effects. There are many ways to interpret AVS and EqP data;

suggested approaches include weight-of-evidence and pass/fail.

I would like to present two examples of how AVS and EqP have been used in evaluating

contaminated sediment.

The first example involves a creek that is exposed to landfill leachate. Chemical analyses

(including AVS, SEM, and TOC) were done first, followed by sediment toxicity tests. The

SEM/AVS ratios were high, always >1 (which indicates there may or may not be

bioavailability), and we really couldn’t say from AVS what was happening. There was no

clear pattern in the bioassays. Survival ranged from 18 to 92 percent. In this case, the AVS

theory was not violated; but the lack of toxicity was not predicted. In the end, the AVS

approach really didn’t help us. We used the AVS tool, it didn’t work in this case but the

theory wasn’t violated.

The second example involves polycyclic aromatic hydrocarbons (PAH)-contaminated

sediments near an aluminum smelter. A wide spectrum of analyses was conducted for this

site: water, sediment, and tissue analyses; laboratory sediment toxicity tests; exposure

studies; bioaccumulation and histopathology with fish, crab, and other species; and benthic

evaluations. The data indicate that PAHs are not bioavailable, even though the

concentrations are very high. The source of PAHs was coal and coke material. TOC values

ranged from 0.1 to over 25 percent. The question becomes, how do you normalize to TOC

with “unrealistically” high TOC values. Can you trust the TOC data? EqP predicted

toxicity, but none was observed in laboratory bioassays.

[A discussion occurred on whether or not EqP and sediment quality

criteria (SQC) based on EqP are meant to be predictive of toxicity. The EPA

proposed SQCs based on the EqP methodology are meant to indicate a

concentration of contaminant below which effects are not expected to be

observed according to the ambient water quality criteria (AWQC); in other

words, a contaminant concentration that is protective of the majority of

the benthos. The EqP methodology, if coupled with LC50 can be predictive

of the same toxic effects for which the LC50 was derived.]

As previously stated, AVS and EqP are tools, but sometimes these tools don’t work. Tools

and screening procedures all have advantages, disadvantages, and assumptions. There is

no “single, perfect, universal tool.”

[A general discussion occurred at this point supporting Dr. Chapman’s

statement: “There is no ‘single, perfect, universal tool.” N o single tool

19
AVS Proceedings

will supply all the necessary data for making decisions at any site.

Different tools supply different information, and the tools to be used at

any particular site depend on the conditions at the site as well as the

questions being asked. It was also pointed out that the screening process is

not a simple yes or no process, it is a general evaluation of existing data,

based on the questions needing to be answered, in order to focus future

expertise and resources on where they will be most beneficial.]

There are many additional issues that must be explored before the AVS and EqP

approaches can be utilized to their fullest extent. These issues include estuarine

conditions—what are the salinity effects on metals; riverine conditions and freshets—what

other variables must be evaluated; routes of exposure—food vs. pore water, and

validation—what is it and what does it involve?

20
AVS Proceedings

21
AVS Proceedings

Certain Chemical Aspects of Heavy Metal-Sediment

Interactions:

The Role of Organic Carbon and the Kinetics of Metal

Sulfide Oxidation

John Mahony

Manhattan College

Riverdale, New York

The purpose of this presentation is to discuss the role of organic carbon in heavy metal

binding in sediments relative to pore water concentration and bioavailability. Our work

arose from the anomaly in sediments that had very high SEM/AVS ratios, with respect to

Cu (i.e., much >1) and yet were not toxic. Without any experimentation, we concluded that

something else was binding the Cu and based upon the available literature, we believed

organic carbon was the binding material. We conducted a series of experiments, and based

on the results of that work, we concluded that organic carbon is not terribly important for

some metals (i.e., Cd), but it is for others (i.e., Cu and Pb). The organic carbon

concentration in sediment invariably explains some of the anomalies in the results obtained

to this point. However, there is no general correlation between AVS and TOC.

To study the role of organic carbon binding in sediments, we designed a system that keeps

total anaerobic conditions throughout (i.e., the sequential batch titration procedure) and

used Cd, Pb, and Cu as our test metals. It is essential to maintain anaerobic conditions

because iron sulfide, which is essentially the AVS, will be converted to hydrous iron oxides,

thereby destroying the ability to make measurements of binding or AVS. The system

consists of a series of flasks containing sediment and de-aerated water. Nitrogen gas is

continuously pumped into the flasks to maintain anaerobic conditions. Metals are added to

each flask at known concentrations, with one flask being the control (i.e., no metal

addition). This system is fairly uniform with equilibrium being reached. We believe we have

a quasi-pore water condition, so that the metal concentration that we measure in the

overlying water is actually what one would find in pore water under similar circumstances.

We also measure AVS in the sediment at the same time that we perform the experiment.

This system prevents sediments from converting any of their chemical properties during their

exposure to metals.

The system was verified (i.e., determined that the metal is actually reacting with the “iron

sulfide” in the sediment) by adding iron sulfide (or sediment) and the metal of concern to

22
AVS Proceedings

each of the vessels, and then measuring the concentration of both iron and metal in the

water. When you reach a molar ratio of 1, the AVS is exhausted (i.e., no more binding sites

for the metal) and the metal is released to the pore water. As metal is added to the system,

iron is released until the point is reached where it levels off because there is no more AVS

and no more iron can be released by adding more metal.

The organic carbon in sediments responsible for binding metals is primarily humic material.

The quality of organic matter does make a difference, and in some cases possibly a very

large difference. It is suggested that, for SQC, organic carbon does play a role in partitioning

metals and, in some cases, can play a dominant role. Binding between the humic material

and metals occurs as the humic material comes in contact with water, the carboxyl and

phenolic groups become protonated, protons are released, and the metal is sorbed.

However, these are not true chemical bonds because the process is absorption, not a true

displacement reaction. The bonding energies here are less than those energies that would be

found in true chemical bonds.

The sorption theory has a number of different aspects to it, and perhaps the simplest of

these is the Langmuir Theory. We have used the Langmuir Theory to understand the

relationship between AVS, organic carbon binding, and metal availability. At low aqueous

concentrations and low sorbed concentrations, the Langmuir Theory can be used to estimate

the partitioning coefficient. Instead of using the sorbed concentration in the usual context

(i.e., mg/kg or moles/g-dry), we use mg/g-organic carbon.

For each of the metals used in the Anaerobic Sequential Batch Titration Procedure we

developed three curves: the first shows a total sorbed metal, the second is non-AVS sorbed

metal, and the third is organic carbon normalized sorbed metal. In addition to what was

reacting with the AVS, you’re plotting the metal that was sorbed divided by the organic

carbon content of the sediment. For Cu, there is a very tight, well-defined trend concerning

the absorption phenomenon, and the trend follows the Langmuir Theory very well. The

trend was not so dramatic for Pb. From our data and use of the Langmuir Theory, we

showed the maximum amount of metal that can be adsorbed on an organic carbon

normalized basis. Organic carbon becomes important for Cu and Pb because there is a large

amount of sorptive capacity based upon the carbon concentration. Metals first bind to sites

associated with AVS and then to organic carbon sites. By knowing the sorptive capacity of

sediments (i.e., the amount of organic carbon present) and the concentration of AVS, the

ability of sediments to remove metals from solution (i.e., from the interstitial pore water)

23
AVS Proceedings

can be determined. This binding action then makes the metals unavailable for biological

uptake.

“Does organic carbon have an affinity for other metals?” is a commonly asked question.

Our next step in understanding metals’ availability is an attempt to answer this question.

Our work has focused on a particular suite of metals because of the electrodes available at

the time to measure metal concentration. As analytical techniques improve our research into

the significance of other metals will continue.

24
AVS Proceedings

25
AVS Proceedings

The Influence of Sediment Stability and Disturbance

on Marine Amphipod Toxicity

Jack Q. Word*

Battelle Marine Laboratory

Sequim, Washington

Jeffrey A. Ward

Battelle Marine Laboratory

Sequim, Washington

Margaret R. Pinza

Battelle Marine Laboratory

Sequim, Washington

Bioassays are routinely used in the Superfund process to identify potential threats to

resources associated with exposure to sediment-bound contamination. It is important that

the potential causes of observed toxicity be identified so the appropriate remedial options

can be considered. We must be able to distinguish between effects due to physical factors

(i.e., grain size) and various types of contamination because different clean up actions will

be applied based on the cause of toxicity (i.e., physical factors vs. non-persistent chemical

compounds that are rapidly lost or changed vs. persistent chemical compounds that tend to

remain associated with the sediment and do not change chemical structures rapidly). The

results of laboratory toxicity tests are used to initiate clean up activities; remediation is

often terminated when the laboratory tests indicate clean up goals have been attained and

the residual contamination will not adversely affect the environment or human health. We,

the scientific community, are good at identifying effects due to exposure to a single

contaminant, as well as predicting what the concentration and effect will be, based upon

equilibrium partitioning theory and AVS. However, we are less certain when it comes to

complex contaminant mixtures.

A fundamental area of research in our Marine Sciences Laboratory (MSL) is the

development of methods to better distinguish between potential causes of the observed

effects. Our approach, a stability/disturbance testing scenario, can be used to implicate

broad classes of potential factors that cause toxicity. I would like to use this platform to

present the theory behind our approach as well as examples of its application.

The stability/disturbance testing scenario involves conducting multiple 10-day toxicity tests

using the burrowing infaunal amphipod, Rhepoxynius abronius. Sediment treatments are

tested following standard procedures (ASTM 1992). The amphipods are added to the test

26
AVS Proceedings

containers after varying time intervals following initial disturbances created during the

preparation of the sediment treatments (up to 100 days). The final exposure occurs after

redisturbing sediment that has undergone the entire period of stability. The change (or lack

of change) in survival compared with the degree of disturbance and periods of stability

provide indications of which sediment characteristics controlled the initial toxicity results.

Differences in test organism sensitivity among various test periods are evaluated by

conducting reference toxicant exposures and assessing amphipod survival in native (control)

sediment. These procedures allow comparisons to be made among tests conducted at

different times.

The primary question examined under the stability and disturbance testing scenario was

whether disturbance of sediment conditions influences toxicity. Tests conducted at MSL

have revealed three basic patterns of response. In the first response pattern, there is no

change to toxicity under any of the conditions of stability or disturbance. In the second and

third response patterns, toxicity decreases during increasingly longer periods that the

sediment is left undisturbed. The difference between the second and third case is that under

the condition of sediment redisturbance toxicity will either increase or decrease relative to

the stable conditions. In Response Pattern 2, the redisturbance toxicity is equal to the final

toxicity observed after periods of stable conditions. In Response Pattern 3, the

redisturbance toxicity increases beyond the toxicity observed during stability and

approaches the toxicity observed during the initial disturbance tests conducted prior to

allowing stable conditions to develop in the laboratory.

Response Pattern 1

When disturbed sediment conditions provide the same level of toxicity as stable sediment

conditions, the contributing factors are proposed to be those sediment attributes that do not

change as a result of disturbance. An example of this type of sediment characteristic is

sediment grain size. The grain size of a sample is unaffected by disturbance, and the

toxicity resulting from less than optimum grain size distributions is the same under both

conditions of sediment disturbance and stability.

Response Pattern 2

When toxicity decreases during periods of stability and does not increase after

redisturbance following long periods of stability, the cause of the observed toxicity is

probably related to a change in the availability of toxicants through time. Examples of

27
AVS Proceedings

contaminants that appear to behave in this manner are: interstitial water ammonia, easily

metabolized toxins, strongly bound contaminants, or volatile materials. The toxic forms of

these materials are lost or sequestered during the sediment stability period, and do not

become bioavailable after disturbance.

Response Pattern 3

When disturbance increases toxicity, the contributing factors are those sediment

contaminants that again become available in toxic form after disturbance. Examples of

contaminants that appear to become more bioavailable during disturbance include nonpolar organic compounds. The disturbance appears to increase the availability of the toxic

forms of these chemicals. One method that might explain this increase in toxicity is that

disturbance increases the dissolved forms of the contaminants that are thought to be the

most toxic. The same process may occur during disturbance of sediments with heavy metals

that are bound to AVS.

[This topic initiated heavy discussion from the audience. It was suggested

that the spiked-sediment tests on which EqP and AVS are based are

realistic and not “worst case”. There was general agreement that spiked

sediments need to be aged before being tested. Furthermore it was stated

that field data compared to laboratory data indicate that the best laboratory

data come from sediments that are disturbed as little as possible, and that

10-d acute tests with R. abronius are reasonably robust at predicting

whether or not non-acute effects will be manifest in the field.]

The experimental procedure and observed response patterns outlined above were used to

help explain experimental results obtained from different studies conducted at the MSL.

Three examples of the use of this testing strategy were evaluated. In the first case,

sediments with no measurable chemical contamination obtained from fine-grained sediments

of Alaskan Glacier Flour origin; coarser but still fine-grained sediment from Oakland

Harbor, California; and coarse-grained sediment from the test organisms native sediment at

West Beach, Washington were evaluated over periods of more than 60 days. In each case,

the survival of the test organisms was consistent throughout the period of exposure during

varying periods of stability or under disturbance conditions. Survival in Glacial Flour

sediment was approximately 68 percent; in the coarser, but still fine-grained sediment from

Oakland Harbor, the survival was 75 percent; and in the coarse-grained native sediment

from West Beach, the survival exceeded 90 percent.

28
AVS Proceedings

In the second response pattern, sediment cores were obtained from hard-packed sediments

that were buried more than 10 feet deep. One of these sediments (the older bay muds of

Richmond Harbor, California) contained no measurable persistent chemical contaminants

(pesticides, PAH, polychlorinated biphenyls (PCB), organotins, etc.); the other sediments

from the deeper more compacted sediments in New York Harbor had a range of chemical

contamination. The survival of test organisms after the initial disturbance and mixing of

sediment was very low for each of these areas. As the time under stable conditions

increased, survival also increased until it was more than 85 percent after periods of more

than 40 days of stability. The final tests under redisturbed conditions showed the same

high survival observed during the latter stable conditions.

Under these examples of Response Pattern 2, the mortality of the amphipods was

attributed to the concentrations of ammonia in interstitial water that were high at the outset

of the experiment, eventually reduced during stable conditions, and remained low under

redisturbance. One explanation for elevated ammonia concentrations during the early parts

of the testing period that were not encountered during the latter periods, is that the nitrogen

cycle and subsequent ammonia concentrations are controlled by bacteria. The populations

of these bacteria were not sufficiently large at the beginning of the experiment because of

their rarity in deeply buried sediments. Once the sediments were under aerobic conditions

the bacteria were able to increase in abundance to population sizes that were able to

denitrify the available ammonia production from decay of organic matter to the less toxic

nitrates and nitrites.

In the third response pattern, the initial disturbance and mixing of sediments resulted in high

toxicity. Under increasingly longer periods of stable conditions the survival increased.

However, under the redisturbance condition (after long periods of stability) survival

decreased again to the level observed during the original disturbance period. Sediments that

provided these types of responses were obtained by performing sediment spiking

experiments using multiple sediment types and various concentrations of radio-labeled

DDT and endrin as contaminants. We also used natural sediments that were contaminated

with PAH obtained in Alaska and another sediment obtained from Oakland Harbor,

California with elevated PAH and tributyltin (TBT). The increase in toxicity under

disturbance situations and the decrease under stable conditions is hypothesized to be a

result of the disturbance of equilibrium conditions between more bioavailable (dissolved)

and less bioavailable phases of the contaminants. The disturbance increases the availability

and exposes the test organisms to more toxic fractions of the contaminants.

29
AVS Proceedings

The stability and disturbance testing procedure outlined above appears to allow

investigators the ability to distinguish between groups of potential causative factors of

toxicity. The changes in toxicity seem to be related to physical and chemical changes that

occur during the periods of disturbance and stability. Our present testing protocol of

disturbance and mixing of sediments just before testing appears to be a method that

maximizes potential toxicity, and as a result, is protective of the environment and therefore

more conservative.

The issue now becomes, what should be done with this type of information? If toxicity is

observed that does not change under stable and disturbed sediment conditions then the

effect is potentially due to factors that are not influenced by disturbance (e.g., sediment

grain size or food quality or quantity). If toxicity is influenced by disturbance and stability,

then the factors controlling the toxicity are probably related to the influence of disturbance

on the bioavailability of the contaminant. This approach to analyzing sediment toxicity is

different from our standard practices, but it may be a method that will allow us to identify

and control for various toxicity effects. The thrust of this research is that the type of factor

causing toxicity should be determined because sediments should be processed and

remediated differently based on the type of contaminant causing the observed responses.

There are other issues that should be addressed when evaluating laboratory bioassay data,

For example, are we underestimating the effects in the environment? A response observed

under laboratory conditions will likely also be seen in the field, but we can not be sure there

will be no effect in the field when no effect is observed in the lab. Also, the effect of

sediment holding times should be reevaluated -- are six months too long? The use of

sediment quality criteria with respect to sediment evaluations, and the ability to predict

effects of contaminant mixtures are among the issues that need further study.

[An extensive discussion occurred at this point, first over other approaches

for determining which sample component is responsible for observed

toxicity (e.g., concurrent elutriate tests and TIE procedures), then over

sample handling procedures. The questions of concern were: How do

sample handling procedures affect toxicity test results? Are existing

protocol handling procedures appropriate? and What are the proper

handling procedures? It was agreed that sample handling is an important

issue but nothing was resolved.]

30
AVS Proceedings

31
AVS Proceedings

KOC and KDOC in Sediment Pore Water

Jim Brannon

U.S. Army Corps of Engineers

Vicksburg, Mississippi

When I was asked to come to this workshop, I thought long and hard about what subjects to

discuss. What topics would really meet the objectives of this workshop? This workshop

deals with how to use EqP and AVS to evaluate the problems of hazardous waste sites. I

have therefore tailored my talk to address this topic. My approach has been to examine the

recent literature on partitioning of non-polar organic compounds to organic carbon and

couple that with work we have been conducting at the Waterways Experiment Station.

Clay particles are coated with iron, manganese, and aluminum oxides; organic matter; and

various other minerals. This system is not as simple as it appears when examining an

idealized representation. Nature is altogether more complex. Rebhun et al. (1992)

examined plating out humics onto clay particles. They found that they could attain

approximately 60 percent coverage of the clay surface, while the other 40 percent remained

uncovered. After that point, the organic matter began stacking on other organic carbon.

This study demonstrates, in addition to our own experience with contaminant sediments

from such places as New Bedford, Indiana, and Black Rock harbors, that the situation is

rarely simple. I like, and use, the partitioning theory because I'm a chemist. I think sulfides

are important and I think organic carbon is important in controlling the concentrations of

metals and non-polar organics, respectively, in pore water. Things that are predictable make

our job easier, but when it comes to actual, real-world sediments, we are often surprised. In

other words, the data do not fit our preconceived notions.

Rebhun et al. (1992) investigated partitioning in a system consisting of 10 percent clay, 90

percent sand, and various concentrations of adsorbed humic acid. Until the concentration

of humic acid reached approximately 0.5 percent, a linear relationship existed between the

partitioning coefficient and the fraction of humic material. Below 0.5 percent humic acid,

behavior became non-linear as other factors, such as partitioning to clay minerals, began to

get more important. At higher organic carbon contents

(> 0.5%), organic carbon is the dominant sorbent, but at lower levels, other sediment phases

become increasingly important. A review of the literature shows that many factors affect

KOC and KDOC. These include pH and/or ionic strength (Schlautman and Morgan 1993),

32
AVS Proceedings

surface reactions (McGinley et al. 1993), and competitive sorption behavior with mineralbound humics (Murphy et al. 1994).

I attended the American Chemical Society (ACS) national meeting in San Diego in March

and McGroddy reported on the work she and Farrington had done (McGroddy and

Farrington 1994). They found that inputs of PAHs in the form of soot affected the

availability of PAHs and lowered their partitioning into pore water.

Contamination in sediments can affect the partitioning behavior of PAHs. Chin and

Gschwend (1992) found that sediment pore water colloidal material from a contaminated

area are better sorbents for PAHs than colloidal material from cleaner areas by a factor of

two. It may be a result of higher lipid content in pore water, higher oil content, or the nonpolar nature of organic matter. Examining the nature of organic matter is a difficult

problem. There are four to five different methods researchers are using to investigate the

polarity of sediment and soil organic matter. Of course, oil contamination in sediments,

something that occurs in just about all man-impacted sediments, will enhance the

absorption of non-polar organic contaminants. From our own experience, anomalous

sorption results occur when coal or slag is present in sediment. These man-generated

components bind non-polar organic contaminants more tightly than do humic acids.

The hydrogen/oxygen (H/O) ratio is one measure of the aromaticity of carbon present in

sediment. The higher the ratio (i.e., the more hydrogen with respect to oxygen) the higher the

aromaticity of the organic carbon. The data of Grathwohl (1990) and Garbarini and Lion

(1986) showed that as the log H/O ratio increases the log KOC also increases. Gauthier et

al. (1987) showed that as the fraction aromatic carbon in humic acids isolated from

sediments increased, KOC values obtained for pyrene increased. These results are what you

would expect, because the higher the fraction of aromatic carbon, the more hydrophobic the

organic carbon becomes, increasing the ability of the carbon to sorb non-polar organic

contaminants. These results indicate that the quality of organic carbon in sediments,

specifically its aromatic character and the kind and extent of human impacts, affect the

KOC of non-polar organic contaminants.

Davis (1993) measured KDOC values for pyrene with humic substances isolated from soils

and water. He did not see a linear relationship between KDOC and the fraction aromatic

carbon. This implies that partitioning of non-polar organic contaminants to sediment

organic carbon may be even more complex. We don't completely understand the situation,

which should lead to caution.

33
AVS Proceedings

Pore water contains contaminants that are truly dissolved and bound to dissolved organic

carbon (DOC). Contaminant concentration measurements in pore water often show high

concentrations, occasionally even higher than solubility limits. This is because the bound

contaminant concentration increases the apparent pore water concentration. To examine

the constancy of KDOC, we tested a wide range of sediments with different concentrations

of TOC. Pore water TOC concentrations from these sediments ranged from 3.1 ppm up to

177 ppm. Pore water was separated from sediment by centrifugation and spiked with

radioactive fluoranthene. The bound and truly dissolved fluoranthene fractions were

separated using Peter Landrum’s reverse phase partitioning method. We then plotted truly

dissolved fluoranthene concentrations versus fluoranthene associated with organic carbon in

pore water. Good linear correlations were obtained for each pore water. The lowest r2 was

approximately 0.85, but the different pore waters showed very different slopes, indicating

very different KDOCs.

Measured KDOCs are from two to four times lower than predicted by partitioning theory.

Therefore, using KOW to predict KOC, and then assuming that KOC equals KDOC, would

introduce significant error into the situation. As we previously observed with KOC, the

source of organic carbon strongly affects partitioning of non-polar organic compounds.

Now, to the bottom line. What does this mean for use of AVS and EqP at hazardous waste

sites. First of all, KOC and KDOC may not be predicted by KOW and may need to be

measured, especially in contaminated sediments. Hazardous waste sites are by definition

contaminated and will pose real problems in predicting contaminant mobility and behavior.

We really have to be careful. AVS and EqP may be used as a screen, but too many sediment

processes that we don't fully understand are occurring to rely on AVS and EqP for criteria.

It is very difficult to make good TOC measurements in some sediments, much less use TOC

as a predictive tool. Factors other than TOC can affect partitioning, especially in sediments

impacted by people.

34
AVS Proceedings

35
AVS Proceedings

Porewater Toxicity Testing:

A Direct Measure of EqP

Robert Scott Carr

National Biological Survey

Corpus Christi, Texas

Over the past seven years, I have been involved in developing, evaluating, and field

validating porewater toxicity tests. Today I would like to describe these tests and

summarize some of our recent field studies. Our work suggests that porewater toxicity

testing provides a direct measure of EqP. A primary advantage of the porewater toxicity

test approach is the ability to conduct tests with very sensitive life stages of sensitive

species, which is not possible with the solid-phase tests that are now commonly used by

regulatory agencies. Species that have been used successfully in porewater testing include

polychaetes (e.g., Dinophilus gyrociliatus), fish larvae (e.g., red drum, Sciaenops ocellatus),

benthic copepod nauplii (e.g., Longipedia americana), and sea urchins (e.g., Arbacia

punctulata). The polychaete Dinophilus is quite small (approximately 1-mm long) and has a

very short life cycle (7 days), which allows us to conduct full life-cycle toxicity tests using a

minimum amount of pore water. The primary endpoint for the polychaete test is egg

production. For the red drum, we use developmental abnormalities as the endpoint. Again,

it is a very short test, with effects usually elicited within 24 hours of exposure. We have

conducted several different tests with the sea urchins, including cytogenetic, fertilization,

and embryological assays.

In the past, we have routinely used a pneumatic porewater extraction device that allows us

to quickly process large numbers of samples to minimize the variability associated with

storage effects. Normally, samples are processed within 24 hours of receipt by our

laboratory. Some recent studies suggest that centrifugation may minimize the loss of some

highly hydrophobic contaminants. With the sea urchin assays, we are able to test a large

number of samples simultaneously, thus minimizing the between-test variability.

We have conducted field validation studies with sediments from Galveston and Tampa

bays. In the Galveston studies, no toxicity was observed with the standard amphipod

tests, but toxicity was observed with the sea urchin fertilization and embryological

development assays. The Tampa Bay sediments also exhibited excellent concordance

between the predicted and observed toxicity for the sea urchin fertilization test but not for

the amphipod test. We believe that the amphipod tests underestimate toxicity because the

36
AVS Proceedings

sediments are manipulated and “flushed” with clean water, replacing the pore water that

has become equilibrated with the sediment-associated contaminants. We believe that this

use of pore water provides a more realistic representation of the contaminants that are

potentially bioavailable to infaunal organisms.

The sediment triad approach was used in a large multidisciplinary study conducted in the

Gulf of Mexico. Several stations were revisited over a period of two years; sediment

chemistry, laboratory toxicity studies, and benthic community structure were measurement

endpoints. The most interesting data were generated for an oil production platform

situated close to the most northerly occurring tropical reef and marine sanctuary. Because

of its proximity to the reef, drilling and production discharges from an oil platform were

shunted to within about 10 meters of the bottom. Laboratory toxicity test results indicated

the urchin embryological development assay was more sensitive than the urchin fertilization

assay for the contaminants at this site. During this effort, we had the opportunity to collect

sediments during the winter, and then again six months later in the summer. Our toxicity

test results were very reproducible and the same spatial pattern in toxicity was observed.

We also subjected selected porewater samples to a suite of toxicity tests. Toxicity tests

conducted with polychaetes, sea urchins, and benthic copepods had a high correlation,

indicating that the urchin assay is a good surrogate for the kinds of organisms that normally

inhabit interstitial water.

Bulk sediment chemistry data for Zn, Cd, Pb, Cu and mercury (Hg) were compared with

threshold effects levels (TELs) and probable effects levels (PELs) of MacDonald (1994) and

the comparable values (i.e., effects range low [ERL] and effects rang medium [ERM]) of

Long et al. (1995). In the Tampa Bay study, there was excellent agreement between the

predicted and observed toxicity with the sea urchin porewater toxicity test, but little or no

concordance with the amphipod solid-phase test. There was also a high degree of

association between the presence of certain contaminants and the toxicity observed with the

sea urchin fertilization assay. Similarly, there was a high degree of association between the

sediment concentrations of several metals and the observed toxicity with the sea urchin

embryological development assay at the highly contaminated oil production platform in the

Gulf of Mexico, where excellent agreement between the predicted and observed toxicity was

also seen. The sea urchin fertilization and embryo development tests with pore water

provide measures of chronic sublethal impacts on sensitive life stages of sensitive organisms.

The PELs, therefore appear to be more predictive of these chronic sublethal impacts than

the more obvious acute lethal effects. Likewise the TELs appear to be protective for chronic

37
AVS Proceedings

sublethal impacts as well, as demonstrated by the high degree of concordance between the

TEL predictions and the observed toxicity with the sea urchin porewater tests.

It is not particularly surprising that there is often little or no agreement between the

predicted toxicity and toxicity observed with the amphipod test. The amphipod test uses

an adult life stage (the least sensitive stage) and mortality as the endpoint. In addition, the

standard amphipod test protocol underestimates the potential toxicity of contaminated

sediments because the pore water, which has taken weeks or perhaps months for the

hydrophobic contaminants to come to equilibrium, is flushed out and replaced with fresh

overlying water before the start of the exposure. The combined insensitivity of the test

species and endpoint and the artificial manipulations of the test sediment prior to exposure,

result in a test that has limited ability to predict the potential chronic impacts of moderately

contaminated sediments.

38
AVS Proceedings

39
AVS Proceedings

Predictions of Cadmium Accumulation

in Benthic Organisms

André Tessier

University of Quebec

Sainte-Foy, Quebec

Canada

It is evident that the development of rational, efficient and economical strategies to manage

the problem of metal contaminated sediments will depend on our ability to predict how

these remedial actions would improve the water quality and affect aquatic organisms. The

prediction of metal accumulation/toxicity in benthic organisms is complicated because these

organisms can accumulate trace metals from various sources: sediment, water, and food.

The bioaccumulation of metals from each of these sources is influenced by numerous factors:

pH, ligand concentration and type, and various sediment components. To further

complicate the picture, a relationship exists between dissolved and particulate metal.

Therefore, predicting the bioaccumulation of metals by benthic organisms will require an

understanding of the biogeochemical processes involved. I would like to illustrate this with

Cd.

During the last few years, I have been involved in the measurement of Cd in several

compartments of the sediment-water system. These compartments include: interstitial

water of the upper (oxic) layer of the sediment, the adjacent several centimeters of the

overlying water, the particulate component of the oxic sediment and the diagenetic iron (Fe)

and manganese (Mn) deposits (Fe and Mn oxides). Cd concentrations were also measured

in various benthic organisms including freshwater bivalves (Anodonta grandis) and

herbivorous insect larvae (Chaoborus punctipenis).

Our study sites were from about 40 lakes exhibiting a wide range of pH and Cd

concentrations. These lakes were distributed over an area of about 350,000 square

kilometers in the provinces of Ontario and Quebec.

First, I would like to show you that the Cd concentrations found in field collected A. grandis,

agree with the free ionic activity model, or to be more correct, do not disagree with it. This

means that the concentration of Cd in the organism is equal to a constant, times the

dissolved Cd+2 concentration in the overlying water. Second, we will look at the

relationship of Cd in the sediment and the overlying water. This will be described by using

sorption reactions with the aid of apparent surface complexation constants. Then we will

40
AVS Proceedings

combine these two types of concepts, the free ion activity model and the surface

complexation model to predict Cd in A. grandis using sediment variables. Third, I will

apply this model to another organism, Chaoborus.

To test the derived free ion activity model relating the Cd level in the organism to the Cd in

the water overlying the sediments, we plotted field data, tissue concentrations from

Anodonta versus free Cd ion concentrations in overlying water collected with peepers. The

results are a linear regression with an r2 of 0.81. While this result is interesting, the free Cd

concentrations are very small, difficult to measure, and may be variable over time.

Therefore we derived a formula relating the Cd level in the organism to the Cd in the

sediment using sorption reactions.

Why do we use sorption reactions? We use them because if you do some thermodynamic

calculations in the oxic zone of the sediment, you cannot find a solid phase that will

precipitate and control the concentration of trace metals. You have to think of other types

of reactions. The one that is probably most logical is the sorption reaction. An oxic

sediment contains few phases that can sorb trace metals. You can think of iron oxide,

manganese oxide, organic matter, and maybe aluminum oxide, but the latter has not been

fully demonstrated.

The free Cd ion concentration ([Cd+2]) can be related by sorption reactions to the sediment

Cd, iron oxide and organic matter concentrations and the pH of the water, all of which can

be readily measured. Using the field data we generated plots of the measured tissue

concentration versus the calculated [Cd+2] from the field sediment characteristics. This gave

us a linear regression with an r2 of 0.82. When we plotted the same tissue data against the

total Cd concentration in the sediment the points were scattered all over the place.

We then applied this approach to data on Chaoborus punctipenis, an insect larvae that

spends the day in the sediment and then migrates into the water column at night to feed on

zooplankton. The resulting plots of Chaoborus Cd to the calculated [Cd+2] showed a very

low r2, presumably due to the low values of Chaoborus Cd found in the acid lakes. The

correlation is much improved if competition is assumed at the biological uptake sites

between H+ and [Cd+2]. Chaoborus Cd then becomes a function of [Cd+2]/[H+]+K, where K

can be considered as an ‘acidity constant’ of the binding sites. This was not a problem for

the bivalve because no bivalves were found in the acid lakes.

Summing up, any model used to predict metal levels in organisms must be field calibrated,

and to be useful, must be able to predict results at locations (lakes) other than those used

41
AVS Proceedings

for the calibration. To be useful the models also must be based on an understanding of the

geochemical and biological processes involved in bioavailability and bioaccumulation of

metals. Such a model would be a very useful tool for management of contaminated

sediments. We have attempted to create such a model here. It is not perfect and more work

needs to be done on it (e.g., the inclusion of the role of Mn oxide). Field and laboratory

work need to be coordinated because what is taking place in the laboratory cannot

necessarily be extrapolated to the field. But, we have made a start.

42
AVS Proceedings

43
AVS Proceedings

Predicting the Toxicity of Metals-contaminated

Sediments using AVS and Interstitial Water

Normalizations: Results from Spiked Sediment

Laboratory 10-day Lethality and 4-month Colonization

Tests

David Hansen

U.S. EPA Environmental Research Laboratory

Narragansett, Rhode Island

The purpose of my presentation is to provide information, developed from sediment spiking

experiments, that will help you see why we believe SEM/AVS ratios and interstitial water

metal concentrations help explain the bioavailability of metals in sediments that can be used

to predict which sediments may or may not be toxic due to metals. A number of people,

other than myself, could be giving this presentation because this has been truly a

collaborative effort. The presentation will be divided into two principle sections: the first

deals with 10-day lethality tests in which interstitial water concentration and SEM/AVS

ratios were measured; the second is a 118-day salt-water colonization experiment

conducted with Cd-spiked sediments.

The lethality test discussions are divided into two parts: the first part deals with the

amphipod toxicity tests that were conducted at the EPA Narragansett, Rhode Island

Laboratory, because all of these tests had essentially the same experimental design. I’ll then

include all the other salt-water toxicity test results from experiments conducted at

Narragansett Laboratory and the Battelle Laboratory in Sequim, Washington, as well as

freshwater test results from the EPA Duluth, Minnesota Laboratory. This will include data

already published by DiToro and others (DiToro et al. 1990, 1992) on Cd, Cd and Ni, and

data produced by Walter Berry, others, and myself at Narragansett on Cu, Pb, Zn, and

mixtures of metals. The mixture’s data will be used to show why the AVS normalization

requires the summation of the molar concentrations of metals.

The experimental design for all tests followed the same general procedure. Metals were

added to vessels containing sediments, stirred and allowed to equilibrate for 4 to 10 days.

A typical experiment contained a control and SEM/AVS ratios of 0.1, 0.3, 1, 3, 10, and

sometimes 30 and 100. Sediments for each treatment were placed in four individual

chambers; two chemistry and two biology replicates. On day zero of the 10-day tests, one

chemistry replicate was sampled to determine AVS, SEM, and total metals. Also on day

44
AVS Proceedings

zero, between 5 and 20 organisms of the appropriate species were placed in the three

remaining replicates. Peepers had been placed in the sediments of these three replicates the

day before testing began. At the end of the tests, the AVS, SEM, total metal, and interstitial

metal concentrations were measured in the chemistry replicate. The biology replicates had

the interstitial water removed from the peepers and stored before chemical analysis. The

sediments were sieved to check for survival of the test organisms.

Typically, there were two to three sediments tested with a range of AVS concentrations for

each of these experiments. These were all flow-through experiments, so we are reasonably

sure that the organisms were responding to the sediment-associated metal or the interstitial

metal but not the metal in overlying water.

Test data were evaluated on the basis that sediments were classified as non-toxic, when

there was less than 24 percent mortality, and toxic when mortality was greater than 24

percent. When percent mortality was plotted against the SEM/AVS ratio, essentially no

toxicity was observed when the ratio was <1.0. When the SEM/AVS ratio was >1.0,

again, these data are only for salt water amphipods test conducted at the Narragansett

Laboratory, most, but not all, of the sediments proved to be toxic.

When the SEM/AVS ratios were plotted against interstitial water metal concentrations and

the SEM/AVS ratio was <1.0 most of the metal in the interstitial water was below our

detection limits. When it was >1.0, interstitial water concentrations generally were high. It

is this release of metals into the interstitial water causing very, very large concentrations,

10s, 100s, or 1,000s of toxic units, that explains why the chemistry in the anaerobic portion

of the sediment is causing toxicity to organisms that live in the aerobic zone.

Next we plotted toxicity against interstitial water toxic units (IWTU). The IWTU is the

interstitial water concentration divided by the metal-specific water only LC50 for the

specific test organism. The results were generally what was expected. Below one IWTU the

sediments displayed less than 50 percent toxicity, and above one IWTU most of the

sediments had greater than 50 percent mortality. Some exceptions, of course, did occur

both above and below one IWTU. Non-toxicity in sediments with IWTUs >1.0 were

probably the result of other factors that are controlling availability in interstitial water

samples.

We also spiked sediments with equi-molar concentrations of Ni, Cu, Zn, and Cd and

plotted the release of these metals into the interstitial water. Based on the metal sulfide

solubility constants for these metals, the expected order of release would be Ni, Zn, Cd, and

45
AVS Proceedings

Cu. This is exactly what was observed using sediments from two different locations. These

results emphasize the utility of a fundamental understanding of the chemistry of the metals

in sediments to help interpret observations from sediment risk assessments.

Now we're going to look at, not only our amphipod data, but all the data for spiked

sediment tests from published and unpublished sources, for freshwater as well as salt water

sediments. The same patterns appear as before. When the SEM/AVS ratios were <1,

essentially no mortality was observed; when ratios were >1.0, not all the sediments were

toxic, but a large percentage of them were. There was little if any difference between the

freshwater and salt water data.

The reason some of the sediments that had relatively high SEM/AVS ratios, but no

observed toxicity, may be biological rather than chemical. In tests conducted with Neanthes

arenaceodentata, a very complex matrix of worm burrows was observed in the sediments at

low SEM/AVS ratios for sediments spiked with Ni. When the ratios exceeded about 5.7,

the burrowing depth was diminished. At higher ratios the worms survived by essentially

remaining on the surface, avoiding the contaminant.

Summing up this collection of data, when the SEM/AVS ratio was 1.0, the sediments were

not toxic 96 percent of the time. When the ratios exceeded 1.0, the sediments were toxic 74

percent of the time. The same kind of numbers occur when IWTU are used for the

prediction (94 percent, 80 percent). If SEM/AVS ratios and IWTUs are combined the

prediction of non-toxic sediments doesn't improve much (97 percent); however, the

prediction of sediments that are toxic improves markedly (88 percent). We believe it is

important not only to look at the SEM/AVS ratios, but interstitial water concentrations as

well. It really helps you to understand whether toxicity in field sediments is associated with

metals.

The final portion of my presentation will examine the results from a chronic toxicity test

that I developed a long time ago at the Gulf Breeze, Florida Laboratory; it happens to be my

favorite toxicity test. It begins with sediment filled aquaria, that over time are colonized by

larvae and other life stages of benthic organisms occurring naturally as plankton in the

unfiltered sea water flowing through the aquaria. In the initial tests using this experimental

design, chemicals were added to the incoming water, but in the last 4 or 5 years there have

been a number of tests conducted with spiked sediments. Typically, there is a control and

three contaminant treatments, with usually 6 to 10 replicates. In the experiments that have

been conducted to date, an average of 7 phyla, 44 species, and about 3,000 individual

46
AVS Proceedings

animals were found per test. These chronic tests last from six weeks to four or five months.

They test the most sensitive life stages of marine organisms because the settling larval stages

are exposed. The test endpoints included: abundance of individual organisms, numbers of

species of different phyla, and diversity. The water exposure experiments are very

sensitive. The no-effect levels have always proved to be less than the water quality criteria;

the observed effect level of these tests has always been above the water quality criteria

water concentration.

In the experiments using Cd spiked sediments that I will describe, there were 12 replicates

per treatment. Four of these replicates were chemistry replicates and eight were biology

replicates. The treatments we used were control and nominal SEM/AVS ratios of 0.1, 0.8

and 3.0. The chemistry treatments were sacrificed on days 14, 28, 56, and 117 of the 118-

day test. Cores were taken of the sediment, which were about 8 centimeters deep, to

determine the time dependent profiles of SEM, AVS, and SEM/AVS ratios. Peepers were

placed in each of the replicates for interstitial water measurements. In the eight biology

replicates, the first sampling event examined the periphyton communities that developed by

about day 80. On day 118, all the biology replicates were sieved to remove the

macrobenthic organisms.

The chemical analyses of homogenized sediments indicated that SEM, AVS, and SEM/AVS

ratios were essentially constant throughout the experiment. However, the pattern of AVS

concentrations with depth reveals sediment AVS profiles similar to those that occur in

natural systems. These AVS profiles developed during the first two to four weeks, after

which they stabilized. Thus, the SEM/AVS ratio in the 0.8 nominal SEM/AVS treatment

was about 1.5 because of AVS oxidation. The profile persisted throughout the experiment

for all the treatments.

By week six we noticed periphyton communities in the three lowest treatments and the

control, but none in the 3.0 treatment until about day 80. When the biological data on

macrobenthic organism abundance were analyzed, there was no difference at all between the

control and the lowest treatment. The 0.8 treatment had reduced numbers of polychaetes

and an unknown species of nematode, but had no effect on overall species abundance. The

3.0 treatment displayed numerous effects, including reduced numbers of annelids and

annelid species, harpacticoid copepods, and nematodes. The average size of the worm

Neries succinea also was reduced, which we believe was not a growth effect but due to a

delay in colonization of these aquaria.

47
AVS Proceedings

When cluster analysis was performed on the species presence/absence data, three clusters

were produced: the control and the 0.1 treatment, the 0.8 treatment, and the 3.0 treatment.

We know, because the same number of species was present in the 0.8 treatment and the

controls, that the presence/absence data demonstrates that the species composition of the

community has shifted. In the highest treatment both kinds and numbers of species were

affected. When we compared the interstitial water Cd concentrations to the criteria

continuous and the criteria maximum concentrations for Cd in water, the control and the

lowest treatment were always below the chronic criterion. In the nominal SEM/AVS 0.8

treatment the measured interstitial Cd concentrations were always above the chronic Cd

criterion and occasionally exceeded the acute criterion. In the 3.0 treatment the interstitial

water concentration exceeded all water quality criteria. Therefore, biological responses

observed agree with those expected given measured SEM/AVS ratios and interstitial water

concentrations.

So I would like to conclude by saying that sediment toxicity data from spiked-sediment

tests indicate that SEM/AVS ratios and IWTU are particularly useful tools in causally

linking organism response to metals concentrations in sediments and that it seems to apply

not only in acute lethality experiments but also to chronic effects.

Question Could you comment on the relatively small fraction of sediments that were toxic

between the SEM/AVS ratios of 1 and 10?

Hansen Exceeding an SEM/AVS ratio of 1.0 may not result in available metal toxicity

because other binding phases exist in sediments that bind metals and some of the

organisms can avoid toxic sediments. Therefore, the prediction of toxicity is

uncertain. Therefore, SQC for metals, if they ultimately are developed, would be

used to identify acceptable, not unacceptable, sediments. John Mahony has been

conducting research to combine AVS measurements and partitioning coefficients

for organic carbon to permit prediction of those sediments likely to be toxic. We

are hoping ultimately to be able to develop sediment criteria predictive of

unacceptable sediments.

[Dr. Hansen recommends that readers should consult the “Briefing Report

to the EPA Science Advisory Board on an Equilibrium Partitioning

Approach to Predicting Metal Bioavailability in Sediments and the

Derivation of Sediment Quality Criteria for Metals” (EPA-822-D-94-002) for

a complete summary of this approach. It is available from the U.S. EPA,

Office of Water Resource Center, at (202) 260-7786.]

48
AVS Proceedings

49
AVS Proceedings

Marathon Battery:

What Were the Questions?

Mark Sprenger

U.S. EPA Environmental Response Team

Edison, New Jersey

The Marathon Battery Site is situated in New York State directly across the Hudson River

from West Point. During the Civil War, the area was a cannon production facility. Later,

part of the property became a Cd-Ni battery production plant, which is the source of

contamination in the marsh and wetland areas. The final remedy and cleanup levels were

difficult to achieve. There were many arduous conversations associated with these

decisions, and we never went back when the work was in progress to redefine what the

questions and goals were. The purpose of today’s discussion is to evaluate, “what were the

questions being asked, and what tools are appropriate for those specific questions?”

There was physical evidence of Cd sulfide in sediments; although a large amount of waste

was discharged directly into the Hudson River and never reached Foundry Cove, the

sediments near the Kimble Ave. sewer outfall contain up to 12 percent dry weight Cd and

Ni. A cleanup level of 100 ppm Cd was used for this area, although some hot spots with

sediment Cd concentrations greater than 500 ppm were left unremediated because of the

high degree of disturbance associated with the physical removal effort. No specific cleanup

level was established for Ni because it was assumed that Cd and Ni co-occurred and that

by remediating for Cd, Ni contaminated sediments would also be removed.

A study was developed to evaluate monitoring options for the effectiveness of remediation.

This plan included x-ray fluorescence measurements, sediment and water sampling for

chemical analyses, benthos enumeration, in situ and laboratory toxicity tests with Hyalella

(however, the in situ tests were not very productive), and bioaccumulation studies. One of

the problems we faced was that the area under investigation was a very dynamic system

that responds to tidal variations. Temperature and dissolved oxygen changed significantly

over the course of 24 hours. One of our concerns was to be sure that the final monitoring

plan accounted for the natural variability. We are frequently faced with having to interpret

a data set that does not truly represent the system; in this case, the data set is fixed. The

system has dynamic aspects. We review the data and think about the system in terms of

that particular data set, and we often forget to go back and think about the system and

review the original questions.

50
AVS Proceedings

We did observe a high variation in Hyalella survival during the in situ exposures. This was

probably due to test chambers silting over and low dissolved oxygen conditions within the

chambers themselves. In some cases all animals were dead in one chamber while there was

no mortality in the next; the data were very inconsistent. We found that approximately

1,000 ppm was the mortality breakpoint for both Ni and Cd. Sediments in East Cove were

not uniform with respect to grain size. The bioaccumulation data are very interesting, with a

decreasing rate of accumulation during the initial time period, followed by an increasing rate

of accumulation. We believe that these results are due to a storm event that caused

sediments to become resuspended, and thus making the metals more bioavailable. The

bioaccumulation data show that in the more contaminated areas, more metals are

accumulated. These data are presented as background information and to get an idea of

what types of analyses were done at this site.

The collected data were used to answer the three major questions asked relative to site

management:

(1) Are the sediments toxic, and if so, at what concentrations?

(2) Are the contaminants biologically active, or, what is the magnitude of food chain

accumulation?

(3) Does the East Foundry Cove pose a threat as a source of Cd and Ni to the Hudson

River system—is there a potential for catastrophic release to the Hudson River

system?

For these particular site management issues, it was not important to know what was going

on with respect to AVS. The third question was an important issue, but unfortunately the

issues of concern were never well stated.

One major question was not asked: Which of the two metals is causing the toxicity? It was

assumed to be Cd. When AVS was used, the results suggested that Ni was responsible for

the observed toxicity, and not Cd. However, for this particular site, it was not critical that

the toxic element be identified, because the two metals co-occurred and would be

remediated simultaneously. It wasn’t important whether it was Ni or Cd that was actually

causing the toxicity relative to the remediation decision that was made. Based on the

questions we asked, using AVS as a screen could have given the wrong or incomplete answer

(i.e., that there isn’t a problem in East Foundry Cove sediments). The conclusion would be

misleading, because that wasn’t the question. One needs to be site-specific when evaluating

contaminated sediments, and site-specific considerations are too complex for generic

statements. One must think through the issues and the process, use the best tools, and be

conscientious that the users of the tools are asking the right questions and applying the most

51
AVS Proceedings

appropriate tools. Asking the correct questions is an extremely important aspect of the

entire process.

[A discussion followed the presentation concerning the utility of AVS and

EqP for identifying areas to be remediated and the difference between a

purely scientific evaluation and a Superfund site evaluation that is

constrained by management and regulatory policy.

John Mahony (he was involved in field and laboratory work at this site)

felt that using AVS as a screen would have indicated “hot spots” for

cleanup but would not have indicated that cleanup was needed in some

areas where cleanup occurred (whether this latter cleanup was needed was

not clarified). However, in this case it did not matter what was causing the

toxicity (the Cd or the Ni).

Walter Berry argued that there are two questions which should be asked,

which are similar to the ACOE’s complaints regarding dredging: (1) Are

the in-place, relatively undisturbed sediments toxic? This question is one

where AVS and EqP can play a role. (2) Are these metals potentially

available to the system? This is a bulk sediment issue, not appropriate for

AVS and EqP (as it is simply a question of whether the metals will be

mobile and flushed into the Hudson River). The answer provided was

that Superfund operates within a risk management/legal decision

process.]

52
AVS Proceedings

53
AVS Proceedings

Use of SEM/AVS Ratios and Pore Water Metal

Concentrations to Predict Metal Bioavailability in

Field Sediments

Gary Ankley

U.S. EPA Environmental Research Laboratory

Narragansett, Rhode Island

Dr. Ankley was unable to attend; his talk was presented by David Hansen.

As with the presentation that I gave yesterday, Gary lists a multitude of contributors to the

data contained in this presentation, including scientists from the EPA laboratories at Duluth

and Narragansett, John Mahony from Manhattan College, scientists from their contractor

located at the University of Wisconsin-Superior, and the co-research with Landis Hare of

the Université du Québec (Hare et al. 1994).

This presentation will focus on the results of laboratory toxicity tests, conducted with field

collected sediments, from both salt water and fresh water sites, and field experiments that

were conducted by the Duluth Laboratory and by Landis Hare, with laboratory-spiked

sediments placed outside in the field.

The first part of the presentation is on field samples, tested in laboratory toxicity tests. The

salt water location samples were Belledune Harbor in Canada, which had high

concentrations of Cd, Pb and Zn; an unnamed salt marsh in Massachusetts that has high

Zn, Cu and a variety of other metal concentrations; and Foundry Cove, which had high Cd

and Ni concentrations. The fresh water sites included: the Keweenaw Waterway and

Steilacoom Lake in Washington, both Cu sites; Turkey Creek, Missouri, a Zn site, and

Foundry Cove, a Cd and Ni site.

The biological methods included 10-day lethality tests with fresh and salt water

amphipods, marine polychaetes (Neanthes sp.), or the freshwater oligochaete Lumbriculus.

Typically these tests had two biological replicates with 10 to 20 organisms per replicate.

All had periodic or flow-through conditions to maximize the exposure by the sediment and

interstitial water rather than the overlying water. Again, there were multiple replicates so

that we could have both day zero and day 10 chemical and biological replicates.

The analytical determinations were: porewater metals, AVS, SEM, and total metals

concentrations in the sediment.

54
AVS Proceedings

When we examine a figure showing salt-water toxicity data from these three field sites, we

see similar plots to those for the spiked-sediment data I presented in my talk yesterday. In

a plot of mortality against the SEM/AVS ratio we rarely see toxicity below a ratio of 1.0,

and above 1.0 sometimes there is and sometimes there isn't any observed toxicity. This lack

of toxicity above a molar ratio of 1.0 occurs because of the presence of other metal binding

phases in sediments. Similarly, mortality is generally lacking in sediments having interstitial

water toxic units less than 0.5. When IWTUs exceed 0.5, toxicity may or may not occur.

This is due to other binding phases within the interstitial water, such as DOC, that can

complex the metals.

[An extended discussion ensued at this point addressing mainly regulatory

and management issues. The key issues were: how to decide what

sediments warranted cleanup based on this data, and the difference

between management decisions and good science. The bottom line was

scientific evaluations and management evaluations are not the same

thing. Good science requires that you use all your available tools to assess

a site, don't just rely on one. If you have metals contamination you

should use both SEM/AVS ratios and IWTUs along with your other tools.

Management decisions must take into account costs and practicality and

thus are based on something less than ideal science.]

I very strongly advocate that you use both SEM/AVS ratios and IWTUs as endpoints to

classify risks from metals in sediments. With the spiked-sediment data our ability to

classify sediments is about the same for SEM/AVS and IWTUs. With the field-sediment

data, use of IWTUs was a bit better for predicting toxicity. If you are going to evaluate a

site having sediment contaminated with metal, you really shouldn't rely on SEM/AVS

analyses alone; you should also be measuring interstitial water concentrations. Either

endpoint is acceptable to identify those sediments that are not likely to have toxicity. But if

you are trying to sort out causality or identify which sediments and metals are toxic, then

you need both endpoints.

Keweenaw Waterway, as you recall, was a Cu site with huge SEM/AVS ratios and

interstitial water toxic units up to about 20. When the SEM/AVS ratios were <1, the

sediment was non-toxic; when >1 10 out of 10 sediments were toxic. When the IWTU were

<1, one of two was not toxic; when >1, 9 of 9 were toxic. In contrast, Steilacoom Lake,

another Cu site, displayed no toxicity both when the SEM/AVS ratio exceeded one (8 of 8

nontoxic) or were <1. The IWTU never exceeded 0.5, even in sediments having SEM/AVS

ratios greater than one.

55
AVS Proceedings

[John Mahony pointed out that when they measured the carbon in

Steilacoom Lake they found it was very high, between 6 and 10 percent

carbon, while the Keweenaw Waterway had virtually no carbon to bind

excess metal.]

Turkey Creek, a Zn site, had patterns of toxicity similar to the Keweenaw results, no

toxicity below an SEM/AVS ratio of one and some toxicity above one. When the IWTUs

were less than 0.5, 3 of 4 were nontoxic, and 1 of 3 was toxic when the IWTU exceeded 0.5.

The Foundry Cove data had SEM/AVS ratios from 0.5 to 189; IWTUs ranged from 0.4 to

110. Overall, 12 of the 16 sediments were toxic to Hyalella; only two were toxic to

Lumbriculus, a species much more tolerant to these metals. When the SEM/AVS ratios were

less than one, 4 of 5 were not toxic; greater than one, 11 of 11 were toxic. Here's an instance

where molar exceedances were a perfect match with observed toxicity. When the IWTU

was less than 0.5, one of one was non-toxic; 12 of 15 were toxic when there was greater

than 0.5 toxic units. These data again bring home the point of the need to quantify both

interstitial metal and SEM/AVS ratios.

When we look at a figure of SEM/AVS ratios versus percent mortality for the pooled data

from the four freshwater sites we see the expected pattern: at ratios <1 no mortality

observed; >1 some of the sediments were toxic and some were not. This absence of toxicity

can be explained by the presence of other sediment binding phases such as organic carbon.

We saw with some of the data that I showed yesterday that avoidance behavior also can

affect the exposure zone of the organism thus reducing observed toxicity.

When we look at a plot of all the data from published and unpublished studies dealing with

both spiked- and field-sediment toxicity tests, we see the same pattern as with the

individual data sets: no toxicity at molar ratios 1, and a concentration-dependent

response as the ratio increases above one. What's important is there aren't any data in the

upper left quadrant of the plot, that is in the portion of the figure below a molar ratio of one

and above 24 percent mortality. If, in your field study, data occurs in this upper-left-hand

region, metals are likely not the cause of the observed mortality. The IWTU plot of all the

data shows the same thing. Thus, both IWTUs and SEM/AVS ratios may be important

indicators of non-metals associated toxicity.

Now we're going to begin talking about two experiments with spiked sediments that were

placed in the field. Landis Hare spiked sediments with Cd and returned them, in trays, to

the lake from which they had been taken. The study, which lasted longer than a year,

56
AVS Proceedings

looked at Cd and AVS concentrations, Cd bioaccumulation, and insect abundance and size.

Based on sediment chemistry the treatments remained separate for the duration of the test.

The insect abundance and size data were less discriminatory, possibly because many of the

organisms leave the sediment and feed in the water column and may not return to the same

tray. The bioaccumulation data showed that, with the species that are not linked intimately

with the sediments, there is no difference in tissue concentrations as treatment

concentrations increase. But for those species that are more tightly coupled to the sediments

when the SEM/AVS ratios were <1, there were no differences or small differences in the

tissue accumulation; once the ratio exceeded 1 there was a concentration-dependent

accumulation.

Gary has a similar study in progress with treatments spanning a range of Zn SEM/AVS

ratios, with horizon-specific chemistry and benthic community structure analyses.

Preliminary results found measured SEM/AVS ratios clumped around one. He

hypothesized that because the trays were placed in the pond at a time of sulfide production

the sulfide reacted with the free Zn in the trays; thereby, tying up the Zn and reducing the Zn

sulfide molar ratios. In the winter, the molar ratios that were supposed to be >1 were, <1.

In the winter when we would have thought there would be oxidation of AVS, that did not

occur in the high nominal Zn concentration sediments. Zn sulfide, which has a very low

oxidation rate, was the only sulfide present. Therefore, SEM/AVS ratios in those

treatments were maintained at a molar ratio of . That was reflected in the measured

porewater concentrations, where at the beginning of the experiment Zn was detected in pore

water, but as the experiment progressed, the Zn concentration decreased to below detection

limits. As a result, there were no observed effects on benthic communities.

The concern with extrapolating any toxicity test or chemistry measurement data to the real

world is have you captured the exposure properly. This is particularly true when the

collected samples have been homogenized. The concern is the microscale associated with

organisms that burrow into the sediment and the structure of the oxic/anoxic zone. By

homogenizing the test sediments, your conclusions based on these results may not be

relevant to the microscale. I think, if you were here yesterday, you would have seen that the

depth profiles of AVS found in some of the toxicity tests that we conducted, did in fact

have microscale structure and likely relevance to the field. There is a lot of structure that

occurs even in a toxicity test as a result of worms or other species burrowing. I believe that

the telling information is the huge interstitial metals concentrations that do occur when the

binding capacity is exceeded. I think they are sufficient to allow a reasonable linkage

between homogenized lab sediments and field sediments. The evidence that I think is the

57
AVS Proceedings

strongest is the similarity in the peeper measurements made in the field at the salt marsh in

Massachusetts and what we observed in the same sediments homogenized in the laboratory

studies. We have folks here arguing that you shouldn't be measuring AVS because of fine

scale structure in the field. So how fine a scale do you want to do this chemistry folks? Do

you want to get down to the millimeter level in some of your sites? I don't think you do. I

think risk assessments are built on the assumption that we're going to miss some risks, and

what we have to do is maximize the accuracy of assessments keeping costs practical. The

little fine scale analyses that should be made in the laboratory and field research may not be

appropriate in large-scale evaluations.

[Dr. Hansen recommends that readers should consult the “Briefing Report to

the EPA Science Advisory Board on an Equilibrium Partitioning Approach to

Predicting Metal Bioavailability in Sediments and the Derivation of Sediment

Quality Criteria for Metals” (EPA-822-D-94-002) for a complete summary of

this approach. It is available from the US EPA, Office of Water Resource

Center, at (202) 260-7786.]

58
AVS Proceedings

59
AVS Proceedings

Sediment Contamination, Toxicity,

and Macrobenthos:

Comparability of Laboratory and Field Observations

Richard C. Swartz

U.S. EPA Environmental Research Laboratory

Newport, Oregon

The focus of today’s talk is the comparability of lab and field observations, concerning

arguments of sediment contamination and sediment toxicity and biology. Specifically, the

issue of field validation of 10-day laboratory toxicity tests conducted with amphipods.

Amphipods have been used broadly in a variety of research and regulatory programs,

including the development and verification of the various kinds of sediment quality

guidelines (e.g., marine sediment criteria based on the EqP approach and the regional

AETs). Because data from laboratory bioassays with amphipods are used so broadly, the

verification of these tests is very important to all of us, particularly regarding the various

strategies that might be employed.

The field verification of laboratory toxicity tests has to be done with two kinds of

applications in mind: first, the field-collected sediment must be evaluated, and second, the

sediment must be experimentally manipulated (i.e., spiked sediments) so that conclusions

can be made.

Three different examples of field verification will be presented. The first, involving a site on

the Palos Verdes Shelf, focuses on a specific contaminant, DDT and its metabolites. The

second, also involving DDT, begins with spatial correlations and then addresses specific

cause-and-effect relationships relative to spiked-sediment toxicity tests. The third deals

with the use of a model to go from observations of toxicity to predictions of toxicity. To

address the question of relevancy of laboratory tests in evaluating ecological impacts

associated with contaminated sediments, both correlative and investigative analyses were

applied to the results of 10-d acute toxicity amphipod tests (Rhepoxynius abronius,

Eohaustorius estuarius, Hyalella azteca, Leptocheirus plumulosus, Corophium spinicorne,

Corophium volutator, Ampelisca abdita) and field collected metrics. Field parameters

measured included amphipod density, infaunal index, species diversity, and species

richness.

60
AVS Proceedings

Los Angeles County discharges hundreds of million of gallons of effluent into the Palos

Verdes Shelf each day. There is a pronounced gradient within the area, particularly for

organic enrichment and a variety of chemicals. We used sediments from this area to

investigate the correlations among the three elements of the triad—benthic community

structure, toxicity, and bulk chemistry. We conducted three levels of verification with the

toxicity test. The simplest was conducting toxicity tests with amphipods and comparing

results to the benthic community—are amphipods found at sites where the sediments are

shown to be toxic in the laboratory? To aid in our evaluation, we used the index developed

by Jack Word and associates from Battelle that ranges from 0 to 100, with values <10

representing highly toxic sites, values around 50 representing altered or disturbed sites, and

values greater than 80 representing unimpacted sites. We found very few resident

amphipods in sediments shown toxic in laboratory studies. We also found sediments that

were clearly non-toxic, according to laboratory test results, which were associated with

disturbed benthic communities (i.e., the index was around 50). These results corroborate

yesterday’s discussion about the relative sensitivity of laboratory tests, and that it is

possible to have sediments test nontoxic in the laboratory but still have disturbed ecological

conditions. This is why it is imperative to have chronic tests as well as sensitive toxicity

tests. Overall, our results indicated that laboratory toxicity was correlated with degraded

benthic communities, degraded amphipod assemblages, and elevated contaminant

concentrations.

To move on to DDT, we examined sediments from three different study areas—the

Lauritzen Channel (part of Richmond Harbor in San Francisco Bay), an open coastal marine

system on the Palos Verdes Shelf, and a freshwater stream system in Huntsville, Alabama.

The fundamental differences among the sites are in the ecology and species of the

amphipods that were used in the sediment toxicity tests. Pooling results from laboratory

toxicity tests for all three sites, we get a clear concentration response pattern — no toxicity

hits at very low concentrations of DDT (normalized to organic carbon) then a concentration

related response as DDT concentrations continue to increase. This is a remarkably good

correlation, particularly since we’re evaluating field-collected data from three different kinds

of habitats. We have data on amphipod abundance for two of the sites (Lauritzen Channel

and Palos Verdes Shelf). In areas were DDT concentrations are low, amphipod

assemblages are abundant and diverse. There are also many areas were DDT

concentrations are low and amphipods are not abundant. These data are not unexpected

because there are many factors other than DDT that control the distribution of amphipods.

61
AVS Proceedings

As the concentration of DDT increases, the abundance of amphipods decreases until you

get into a quadrant where amphipods are rare or totally absent.

I would like to now focus on the Lauritzen Channel study. Several contaminants of concern

were identified, with DDTs, PAHs, and metals identified as the principal contaminants.

We used a toxic unit (TU) approach to evaluate the toxic contribution made by each

principal contaminant of concern. For metals, the TU was <<0.2 (based on the SEM/AVS

ratio), indicating that metals are not a problem at this site. For PAHs, the TU was around

0.28, which is in the range that toxic effects could be observed, but it is not sufficient by

itself to explain the 100 percent mortality observed in laboratory toxicity tests. The TU for

DDT was >5.0. This assessment suggests that DDT is the primary contaminant causing

observed mortality.

The Huntsville, Alabama, data were taken one step further. We calculated an LC10, which

we believe indicates a threshold level of toxicity of DDT based on spiked-sediment toxicity

testing. The data from this study support our hypothesis that it is possible to obtain

estimates of chemical concentrations that cause effects and/or degradation in the field using

spiked-sediment toxicity tests, and the spiked sediment toxicity test is relevant to sediment

toxicity in the real world. Sediment toxicity tests with sensitive amphipods applied to

field-collected and experimentally spiked sediments provide ecologically relevant

information. This information is particularly important since, with the exception of

screening level concentrations (SLC), all other methods of deriving sediment quality

chemical-specific numbers depend heavily on data from these amphipods.

And finally, a look at the PAH model developed to predict toxicity. There are papers

written that address the sources of error and limitations associated with the model, and the

model does need to be evaluated by the research community. Nevertheless, the model

suggests the following:

• If mortality is < 13 percent, then sediments are not toxic.

• If mortality is > 24 percent, then sediments are toxic.

• If mortality is between 13 and 24 percent, the level of toxicity cannot be predicted.

The model begins with field chemistry and uses equilibrium conditions to convert the bulk

chemical data to make predictions of interstitial water concentrations. The model then

predicts the 10-day interstitial water LC50 for each of the 13 PAH compounds. The model

then calculates toxicity units for each of the PAH compounds, and finally summarizes the

62
AVS Proceedings

individual toxicity units to derive a single toxicity unit for the sediment under evaluation.

The response model is used to predict the probability of significant acute toxicity associated

with those sediments. Sediment PAH concentrations and TOC values were used as input

data for this model, and predicted and observed toxicity were used to verify the model. A

close association between observed and predicted toxicity was obtained, with no significant

difference between observed and predicted toxicity. The model accurately predicted the

probability of toxicity at field sites where PAHs were the dominant contaminant. However,

if PAHs are not the dominant contaminant, this particular model cannot predict the toxicity

due to DDT or any other contaminant, but it can give an estimate of the relative contribution

of PAHs to overall toxicity. If the relative contribution is very small, then you can believe

that something else is the major cause of the observed effect.

With respect to amphipods, there are differences in sensitivity among species and it is

difficult to make generalizations, but generally all the seven species listed above would be

considered “sensitive species.” One species in particular that is not considered sensitive is

Grandidierella japonica. This species has been found to inhabit sediments determined highly

toxic to other species in laboratory toxicity tests. Current research projects are being used

to determine what makes this species less sensitive. Results of our work indicate: (1)

adaptation and/or selection for resistant populations, or (2) in some cases, populations

may be living in a thin, less contaminated layer. G. japonica is a shallow burying amphipod

and could be portraying avoidance behavior similar to that mentioned by Dave Hansen for

polychaetes and by other researchers investigating the effects of creosote-contaminated

sediments on benthos in Puget Sound.

In the grand conclusion of all three examples, I would hope that we would have consensus

on the fact that sediment toxicity tests with sensitive amphipods applied to field collected

and experimentally spiked sediments provide toxicity estimates that are relevant to benthic

ecosystem contamination and degradation in the field.

In response to the question of whether or not our triad evaluation approach to risk

assessment precluded the need for using EqP, I think that EqP is an important part of the

evaluation because we basically made the comparison with the organic carbon normalized

concentrations. That’s how we could find the exposure. I don’t see it at all as an

independent stand-alone toxicity type evaluation. I think that it is critically important and I

would think that we might agree on that also, that we look at these other approaches

concurrently to make the assessment.

63
AVS Proceedings

64
AVS Proceedings

Comparison Between the Results of Toxicity Testing

with Marine Sediments and

Predicted Effects Based on Chemistry

Jack Anderson

Columbia Analytical Services

Carlsbad, California

Columbia Analytical Services (CAS) has developed a new sediment assay, the Cytochrome

P450 Reporter Gene System (RGS), that is based on luminescent biochemistry. Firefly

chromosome material has been incorporated into a line of genetically engineered human liver

cancer cells, permitting application of luciferin/luciferase analyses. San Diego Bay

sediments were simultaneously subjected to a series of chemical and biological analyses.

The purpose of the study was to provide an overall assessment of San Diego Bay near the

naval station. The 10-day amphipod (Rhepoxynius abronius) test was used with solid phase

material, MicrotoxTM and the P450 RGS test were used with organic extracts of sediments,

and oyster larvae and MicrotoxTM were tested with sediment elutriates. Elutriates would

contain both soluble metals and organics, while the organic extracts contained primarily

PAHs.

Reference station sediments were primarily sand; otherwise grain size was quite variable.

There was some concern with the effect of grain size on amphipod survival, but the effects

observed did not correlate with grain size. TOC ranged from 0.1 to 2.8 percent. In

decreasing order of importance, the toxic metals identified were Zn, Cu, Pb, and chromium

(Cr). SEM/AVS analyses were done on all sediment samples. The toxic metals added

together resulted in high levels for the stations closest to the shore; the reference stations had

very low levels. However, the ratio of SEM to AVS resulted in somewhat anomalous

results. Stations away from shore (toward the center of the channel) exhibited the highest

ratios of SEM to AVS; because the sulfide levels were low, 22 of 30 stations had ratios >1.

The PAH data demonstrated a decreasing gradient with distance from shore for all

transects, except the reference transect. The highest PAH concentration, approximately 20

ppm, was measured in sediments collected from the shore stations near the docking

facilities.

The elutriate tests with oyster larvae were not very informative. Two of the larger hits

occurred at the reference stations; stations where PAH and Zn concentrations were highest

did not show significant effects. The oyster larvae data did not correlate with the ratios of

65
AVS Proceedings

SEM to AVS, nor with the concentrations of sediment PAH. None of the 30 elutriates

significantly reduced the luminescence of the bacteria in the MicrotoxTM test. When the

organic extracts of sediments were applied to the MicrotoxTM test, there were some

significant effects, but these did not have a high correlation with sediment PAH

concentration. In contrast, the new P450 RGS results from exposure to the same organic

extracts were highly correlated with sediment PAH content. The solid-phase tests with

amphipods were apparently not sensitive enough as no sediment produced 50 percent

mortality. There were several sediments that produced between 30 and 40 percent

mortality, but these responses did not correlate with sediment chemistry.

The best correlations were obtained for the sediment PAH and P450 data. The observed

effects were associated with PAHs; the concentrations of pesticides and PCBs in sediments

were below detection limits. The PAH and metals data were compared to the EPA

sediment criteria for three PAHs, the Washington State Sediment Criteria and the most

current ERM values. At only one station did all four indicators show an adverse effect.

Two of the stations contained concentrations of three to four metals that were above the

ERM and some individual PAHs above the ERM and the Washington State Sediment

Criteria. Also, the total of high molecular weight PAHs was above the ERM. Therefore, the

ERM values, the Washington State Criteria and the P450 RGS response predicted the effects

observed at the most contaminated stations.

Twenty-two of the thirty stations had SEM/AVS ratios >1, and as per theory, the toxicity

of these sediments was not predictable. Eight sediment samples caused toxicity with

amphipods (i.e., mortality exceeded 30 percent), and for three of these samples, the

SEM/AVS ratios were <1, which suggests no toxicity. The observed toxicity was attributed

to PAHs. Because of the manipulations and oxygenation of the sediments, AVS should

probably not be considered when doing elutriate tests. TOC normalization was not useful

in this case, possibly because there was not a great deal of difference. However, it is an

important measurement and should always be made. We expected the organisms to be

more sensitive than they were. For this particular study, the chemical benchmarks were

more predictive than the laboratory bioassays with oyster larvae or amphipods. The P450

assay is a tool that can be used as a biological test and a detection system. It is sensitive to

dioxin, PCBs, and PAHs. This work was done with sediment extracts; future studies will

include the evaluation of sediment pore waters.

66
AVS Proceedings

Assessing Toxicity in Multi-Metal Contaminated

Sediment:

Application of Theoretical and

Investigative Approaches

Michael Johns*

E.V.S. Environmental Consultants

Seattle, Washington

Frank S. Dillon

E.V.S. Environmental Consultants

Seattle, Washington

We used a suite of tools during a sediment investigation for a site located in an urban

industrial area, adjacent to a freshwater harbor. Sulfuric and hydrochloric acids were

produced at the site, with the primary waste products being pyrite cinders. Although there

are several contaminants associated with pyrite cinders, the contaminants of concern were

Cd, Cu, Hg, and Zn. This site was a good candidate for application of the SEM/AVS

technique because of the contaminants of concern, the presence of iron, and the dynamics of

the system that relates them.

Prior to our involvement, bulk chemical measurements were the only data collected for this

particular site. There is a noticeable absence of organic compounds in the sediments.

Concentrations of Cd, Cu, Hg, and Zn were up to 15,000 mg/kg-dry weight, which

exceeded nearly all screening and/or advisory criteria (e.g., ERLs, ERMs). Based on the

chemical measurements the sediments adjacent to the site were characterized as highly

toxic. As with the other approaches we’ve been discussing, bulk chemistry is a screening

tool and should be used as such. It is a way to characterize the site during the initial

investigation phase and serves a particular purpose—it provides a range, or at least an

idea, of the contaminants that might be present. Bulk chemistry data can also be used to

determine which tool to take from our environmental tool box. In addition, bulk chemistry

data can also indicate whether you are dealing with extremely hazardous wastes or

dangerous wastes that might lead to emergency removal. However, caution should be taken

when making assumptions regarding toxicity based solely on bulk chemistry data. The bulk

chemistry data do provide the rationale and information we need to set up the assessment

approach to determine the degree of toxicity, but further interpretations of the potential

effects of observed contamination should not be done until additional assessment data are

available.

67
AVS Proceedings

We were brought into this investigation because the initial analyses indicated the sediments

to be extremely contaminated and, based on the sediment screening guidelines, the entire

area would need remediation. This particular site is situated in a sector of the country

where criteria are part of the decision process for cleanup, and by using the sediment criteria

for these four metals, nearly all the sediments adjacent to the site required cleanup. Our

assessment process incorporated a suite of tools, including physical and chemical analysis

of sediments, derivatization of SEM/AVS ratios, laboratory toxicity testing, and modified

toxicity identification evaluations (TIEs). The physical data allowed us to better

understand the sediment matrix being tested, which was particularly important information

when evaluating the toxicity test data. Grain size and other sediment characteristics can be

significant factors in toxicity test results and we needed to separate the effect of physical

factors from contaminant exposure.

[A question was brought up regarding the validity of using AVS to

evaluate the potential hazards from sediment Hg contamination. Dr.

Johns felt that because Hg forms a sulfide that is less soluble than iron

sulfide, AVS is appropriate for evaluating the potential acute toxicity of Hg

to benthic organisms, but a separate evaluation would have to be done to

determine the risk of biomagnification and toxicity to higher trophic level

organisms posed by Hg contamination.]

The concentration of chemicals measured during our investigation were quite similar to those

measured during the initial investigation; there had been no major contaminant

concentration shifts at the site over the past few years. SEM/AVS ratios for five stations

were <1 (i.e., no toxicity) and 16 had ratios >1 (i.e., potential toxicity). We could have

stopped the investigation at this point with the conclusion that the majority of the sediment

at the site was toxic. However, SEM/AVS ratios >1 are prone to false positives because of

the other binding factors that may be present in the sediment that further reduce

bioavailability. The highest ratio obtained was 125; the majority of ratios were between 2

and 10. These 16 sites that had ratios >1 were of particular interest to us with respect to

laboratory toxicity testing. Based on the SEM/AVS theory, sediments from the stations

with ratios >1 had more available metals, and hence the potential for toxicity. However, we

didn’t have a lot of confidence in our ability to predict which sediments are going to be

toxic; therefore, it was necessary to conduct the laboratory toxicity tests.

No significant effects were found with the Chironomus growth test. However, results of the

10-d survival tests with both Hyalella and Chironomus led to a modified TIE (i.e., 48-h acute

toxicity tests with Ceriodaphnia)—the Hyalella test had nine hits, the Chironomus test four.

68
AVS Proceedings

The Chironomus tests agreed with no toxicity for the five stations where SEM/AVS ratios

were <1. The Hyalella results agreed less with the SEM/AVS data—in two cases toxicity

was encountered when the SEM/AVS ratio was <1. The Hyalella results suggest toxicity is

not limited to metals contamination, even in a system where metals dominate the

contaminants of concern. SEM/AVS >1 for 16 stations were not predictable with respect to

toxicity, as per theory. Interestingly, the TIE for stations that were toxic in solid phase tests

did not show a match—only one of the stations toxic in the solid phase was also toxic to

Ceriodaphnia in porewater assessments. There doesn’t appear to be a good explanation for

this discrepancy (i.e., toxicity in the whole sediment tests but not in the porewater tests). If

you accept that TIEs have the ability to determine and identify the source of toxicity

observed in whole sediment, then you would conclude that the observed toxicity at only one

station is due to metals—toxicity at the other 12 stations is due to some other contaminant

or physical factor.

[A discussion occurred at this point concerning the potential for changes

in sediment and pore water sample toxicity as a result of handling and

storage. The bottom line appeared to be that handling and storage times

should be minimized; and to better interpret the results of toxicity tests,

the physico-chemical properties of the samples should be measured at the

time of the toxicity tests.]

A qualitative analysis of the benthic community suggests the area is fairly healthy. It is

dominated by vegetative macrophytes, being quite lush up and down the entire system.

Amphipods, as well as aquatic snails, and mussels, were present at all sites. It would have

been very useful to have quantitative benthic data to aid in interpretation of the chemistry

and laboratory toxicity data (i.e., sediment quality triad).

The summary data suggests that by using more techniques we are better able to evaluate the

area of concern. On a site-wide basis, there is gross metals contamination; there are a

certain number of stations for which the SEM/AVS ratios were <1, indicating there should

be no toxicity because the metals were bound and not biologically active. Of the stations for

which the SEM/AVS ratios were >1, only about half elicited toxic responses in laboratory

bioassays. By stopping the investigation before conducting the TIEs, the overall conclusion

would be that approximately half of the area requires remediation. However, the TIE data

indicate that a much smaller area may contain toxic concentrations of metals. In addition, it

is possible to have toxicity in a metals-dominated system even though the SEM/AVS ratio

is <1. This situation is one where additional contaminants (i.e., organic compounds) are

present and, either alone or in conjunction with the bioavailable metals, are causing toxicity.

69
AVS Proceedings

The nature of this study and others involving complex aquatic systems, are the basis for the

question “is there a suite of tools we should always use?” Because of this complexity, we

need to use an array of assessment tools in the evaluation process.

70
AVS Proceedings

Strengths and Weaknesses of the EqP Approach as

Applied to Hazardous Waste Sites Risk Assessment

Spyros Pavlou*

Foster Wheeler Environmental Corporation

Stuart, Florida

John Toll

Foster Wheeler Environmental Corporation

Stuart, Florida

The EqP approach is based on good science. There is sufficient and convincing evidence

that the theory has been validated in the laboratory and the field. EqP is applicable to the

benthic environment in terms of predicting risk. However, this approach is insufficient. EqP

is not the weak link in exposure estimation; the weak link is sediment-water column mass

balancing and bioaccumulation modeling.

How should risks be predicted for higher trophic level species? We should design models

that look for specific parameters and processes that have the largest impact on the risk

estimation. It is not sufficient to use just EqP or AVS in the decision process for site

remediation. Other information, such as the receptors requiring protection, their mobility

and behaviors, the spatial distribution of contaminants and resources within the system,

and the end use of the site, must come into play. The available data should be used to

determine what is accumulated in the prey tissue and what is being consumed by the

predator. In some cases, it makes sense to measure biota tissue concentrations or doses

directly, rather than depending on predictions. This is particularly true for the mid-trophic

level species where it is easy to acquire specimens and analyze tissues. Empirical

relationships have been developed that can be used to predict bioaccumulation. As

indicated earlier, as one moves beyond benthic trophic levels, the influence of the porewater

concentrations is not a big factor in estimating exposure and, therefore, risk.

[An example of the bioaccumulation model developed for the Rocky Mountain Arsenal was

presented.] There are numerous factors (i.e., lake morphology, in- and out-flow rates,

suspended solids, sediment loading) that had to be considered in estimating exposure at

this site. The time required to identify these factors, as well as the associated uncertainty,

justify the measurements of prey items and the use of a direct model instead of modeling

exposure as a function of sediment contamination levels. The EqP approach is valid, but

the way it’s being used creates potential problems, particularly if you don’t know how to

define the important variables. Sensitivity and uncertainty analyses should be used to

71
AVS Proceedings

identify these parameters. This will give flexibility to the management decision regarding the

end effect.

Sensitivity analysis allows the variables to be ranked in terms of their relative contribution

to the overall uncertainty about risk. The results of this analysis can be used to identify

which parameters are important enough to be modeled probabilistically, and to develop a

cost effective modeling strategy that improves the data. By using all the analytical tools

and considering all the options presented by the data available for a site-specific risk

assessment, good customized procedures can be developed to predict contaminant

bioaccumulation and risk in upper trophic level species.

To answer the questions being asked about hazardous waste sites, we must use all the

available tools and rationalize the differences between tools to form a defensible approach.

What is the best approach for predicting risk to upper trophic level biota? The first step is

to design models and data collection efforts with the goal of maximizing expected value of

information. Upper trophic level bioaccumulation should be modeled probabilistically (less

information is required as compared to point estimates, allows for sensitivity and

uncertainty analyses to identify driving parameters, provides more useful information for

risk management decisions). The shortcomings (i.e., reliability of parameter distributions,

correlation assessments) of Monte Carlo analysis must be recognized and site-specific data

should be used to calibrate models. Finally, use decision analysis to balance risks, costs,

and benefits of risk management options.

The applicability of parameter values from the scientific literature should be closely

scrutinized. For example, biomagnification factors generally are not transferable to other

sites because of differences in methods for estimating bulk concentrations in exposure

media, differences in the many physico-chemical processes that govern the rate of uptake of

exposure medium contaminants by biota, and differences in the biology and ecology of sites.

Therefore site-specific biomagnification factors should be used. Site-specific data improve

the reliability of parameter distributions by incorporating more data. Field data should be

used to modify data to improve predictions.

One must consider all the options in designing models and data collection efforts, including

mechanistic models (e.g., EqP), semi-empirical models (i.e., food webs), empirical models

(e.g., statistical estimators of contaminant biomagnification factors), and the use of direct,

site-specific data to improve reliability and investigate parameter correlations (e.g., sample

prey tissue rather than predictions from sediment data). It is important to use empirical

72
AVS Proceedings

relationships. Because predictive bioaccumulation models inherently have a high degree of

error, a direct approach is recommended to measure the concentration of contaminants in

prey items.

In conclusion, EqP is a good approach for benthic ecology, but it must be used in

combination with other tools to establish risk-based cleanup goals. Models should be built

to answer specific questions. Sharpened, well-rounded analytical expertise and creativity

are needed to take advantage of emerging [and present] tools.

73
AVS Proceedings

74
AVS Proceedings

Panel I Discussions

Chris Ingersoll

National Biological Survey

Brenda Lasorsa

Battelle Marine Science Laboratory

Sequim, Washington

John Scott

Science Applications International Corporation

Question 1: What can and cannot the SEM/AVS ratio really tell us?

The following consensus statement was drafted during this panel discussion

and later refined as more information became available and agreement was

reached.

For sediment in which the contaminant(s) of concern include one or more of

the following: Cd, Cu, Ni, Pb, and Zn, an SEM/AVS molar ratio of <1 indicates

the bulk of these metals are complexed with sulfides and are essentially not

bioavailable and hence acute toxicity from these metals is highly unlikely. An

SEM/AVS molar ratio >1 does not necessarily indicate that these metals are

bioavailable and acutely toxic; other factors may be affecting their bioavailability

and toxicity. Toxicity refers to laboratory tests only and does not necessarily apply

to bioaccumulation of the metals or alterations in the benthos. {Noted Cautions:

(1) The numerator and denominator must be considered when evaluating

the ratio, almost no AVS compared to some SEM has a different

meaning than a similar ratio wherein the AVS concentration is much

higher; also, very low concentrations of both SEM and AVS should not

be considered predictive.

(2) The consensus statement is only valid if AVS concentrations are above

operational detection limits.}

The panel indicated that SEM/AVS allows you to look at availability of metals for

interaction with organisms, if the ratio is <1 then metals are not available, if >1 then metals

may be available. Toxicity is not predicted if the ratio is >1. Situations where biology does

not agree with chemistry will occur when you try to use chemical models to predict

biological effects.

John Mahony noted that if the Cr source is chromate, AVS is applicable; if chromate is not

the source, this approach should not be used.

75
AVS Proceedings

Dominic DiToro indicated that Hg will not be amenable to SEM/AVS analysis; the issue is

methylation, not ionic Hg, so the causative mechanism (of adverse effects) is different from

other metals. SEM/AVS has not yet been subject to uncertainty analyses; this needs to be

done but until it is, the absolute value of 1.0 is all that can be used. He further clarified

that, for site investigations, it is generally not enough to just determine the SEM/AVS ratio;

toxicity tests are often also necessary and a highly valuable tool. Certainly, remediation

decisions require more than just chemical measurements.

Alyce Fritz felt that if the SEM/AVS information was available, then some decisions not to

proceed will be based solely on the chemistry data. However, if there is a choice of whether

to do SEM/AVS or toxicity testing, the preference will be for toxicity testing. In an ideal

situation, you should get both pieces of information. The controversy over whether or not

SEM/AVS can stand on its own as a decision-making tool for Superfund will most likely

continue. How should this tool be used?

Jack Word reminded us that the SEM/AVS approach is indicative of acute toxicity and is

not applicable for estimating chronic or sublethal toxicity. The value of this approach is

that it allows one to identify that metals may be the issue, helps identify the contaminants

of concern, and points you in a direction for further study. However, Dominic DiToro

countered saying some chronic data is available. Chris Ingersoll indicated that it is

necessary to look at bioaccumulation. The SEM/AVS ratio holds, for the most part, for

bioaccumulation; however, there are exceptions to this approach also. Data were presented

where midges accumulated Cu with an SEM/AVS ratio <1. Dominic DiToro further stated

that at ratios <1, in a pure system there are virtually no metals available for interactions

under any circumstances in biological organisms, which is the way the theory is presented.

At ratios >1, depending upon whether you’re addressing the real world or the ideal system,

there are metals present that may or may not be able to interact with biological organism.

The consequences of the interaction may be expressed in acute bioassays or in the benthic

community structure. Metals are not available at ratios <1; metals may be available at

ratios >1, the consequences of which are determined with bioassays. The area of

bioaccumulation is one that requires additional investigation. It is uncertain if good data

sets concerning SEM/AVS versus bioaccumulation exist. There are some definite trends for

these two properties, but there are also inconsistencies. There is an additional need to look

at bioaccumulation as one of the subtle responses.

It was suggested that tools other than SEM/AVS be considered (aside from toxicity tests

and community structure analyses), including porewater toxicity units (e.g., which address

76
AVS Proceedings

free metals); TIE; other physical/chemical parameters (e.g., TOC, grain size, ammonia, other

confounding compounds {multiple metals, organics}); sample collection, handling, and

analytical techniques.

Also, the uncertainty boundaries must be established for the ratio, and there is a need to

qualify whether bound metals will ever be available again. There was considerable

discussion regarding the lack of toxicity when ratios are >1. The finding of no toxicity may

be species dependent, with some species actively avoiding toxic sediments.

The issue of acceptable mortality loss was raised. As per recommended protocols, the

sample must show a statistically significant difference from control. In general 20 to 25

percent mortality has been accepted for the cutoff for significant mortality. How does

mortality close to cutoff reflect on the possibility of an organic toxin also? When mortality

data approach the cutoff, it was suggested to consider using an additional tool.

Dominic DiToro closed this question with the comment that we are really asking two

questions: what does the SEM/AVS ratio tell us with respect to the field and laboratory

data, and how should we initiate a site assessment. These are two separate questions, with

the second based more on practicality.

Question 2: How do temporal (e.g., seasonal) and spatial changes in physico-chemical

parameters affect our assessment of hazardous waste site conditions?

Spatial and temporal effects are two very different issues. There are two spatial scales we

must be concerned with: the macroscale (i.e., the area that is the object of the investigation)

and the microscale (i.e., the area that the biota is exposed to). The macroscale is clearly

objective(s) driven; it is a function of the objectives of the investigation, the site, the

characteristics, etc. Everything that is done for the investigation (i.e., AVS/SEM, EqP,

toxicity tests) applies on a macro-scale level. The microscale, the area immediately

surrounding the organism, is affected by the presence of biota; in fact, they are able to

oxidize metals right around their tubes or in the immediate vicinity. However, there is little

evidence that the microscale effect has any impact on the development of the data (on a

macroscale level). The present macroscale approach seems to be working. The microscale is

probably not too important with respect to toxicity and bioaccumulation, although it may

explain some of the anomalous bioaccumulation results. The effects of the microscale on

toxicity or bioaccumulation remain an area for active research.

With regard to the temporal situation, seasonal changes may well occur, and certainly do in

some systems (but perhaps not in others; it is probably too early in the research cycle to

77
AVS Proceedings

make any predictions as to why, when, and where this may occur). There is evidence that

freshets in rivers, and the conditions associated with additional sediment deposition, have

a major effect (e.g., possibly really reducing AVS concentrations). However, the point was

made that the “noise” may be too great to see real differences (“noise” would include

recognized sediment heterogeneity) and, very limited sampling occurs in the winter months,

even after storm events. This represents an important data gap that should be addressed.

A specific question was asked regarding diurnal fluctuations of SEM/AVS in estuaries and

creeks? The general feeling was there was probably not much fluctuation.

Another issue addressed was how to deal with isolated pockets of contaminants (e.g.,

creosote). When assayed in the lab, they show a much more consistent (and possibly lethal)

distribution.

Overall, it may be very difficult to decide in the field if the measured differences are from

spatial or temporal effects.

Question 3: Which parameters of the system should be measured and how can these

data be applied to an entire ecosystem?

This question was not discussed due to lack of time. On a system level, we can’t think

much beyond bioaccumulation and potential biomagnification.

78
AVS Proceedings

Panel II Discussions

Ted DeWitt

Battelle MSL

Jim Meador

NOAA, NMFS

Philippe Ross

The Citadel

Question 1: Is equilibrium a good assumption in the environment?

The Panel broke this question into four parts:

(1) What is equilibrium?

(2) What factors/processes disrupt equilibrium?

(3) On what spatial and temporal scale do these processes occur?

(4) When non-equilibrium conditions occur, does it matter?

The Panel agreed that to answering Question 1, the first step was to define equilibrium and

provided the following definition: the condition when a contaminant concentration in the

various phases does not change over a significant period (months to years), assuming that

the total contaminant mass remains constant.

It was pointed out by Peter Landrum that this is the definition of steady state, not of

equilibrium. He defined equilibrium as the condition when the free energy in a system is a

minimum. Dominic DiToro agreed in definitive terms, but countered that we’re working on a

smaller scale (i.e., equilibrium between sediments and interstitial water), and that the

definition provided by the Panel is probably a good working definition.

Two problems arose regarding the equilibrium:

(1) Equilibrium is not really measurable in sediments.

(2) Equilibrium is only partly attained. What is meant by EqP is a metastable

equilibrium (i.e., no change) between sediment and pore water.

On this basis, equilibrium can operationally be defined by the definition provided by the

Panel, though technically speaking, this really defines steady state.

79
AVS Proceedings

To address the second and third parts of the equilibrium question, what factors/processes

disrupt equilibrium and on what spatial and temporal scales do these processes occur, the

Panel developed a table (see next page) that addressed the various relevant processes that

could violate the assumption of steady state in the sediment. They attempted to estimate

the frequency that these processes occur in nature, over what areal scale, and what impact

they would have on the bioavailable concentration of the metal. The table was offered as a

strawman in hopes of stimulating discussion regarding the equilibrium processes. The

audience requested a checklist that could be used by investigators to make decisions

regarding the equilibrium models. It’s the heart of the assumptions of equilibrium that

cannot be readily accepted. It’s the heart of the model that we have no information on for

most of the sites. Something is needed so information can be collected to make a

determination whether equilibrium exists, or does not exist. If these models are going to be

pushed, the basic assumptions of the models must be checked as to whether they exist,

they’re not intuitive at the Superfund sites. Are the assumptions of equilibrium generally

met, or generally not met, at a site? This is the basic question.

There was also a request for information as to when the assumptions will be violated. It is

imperative that site managers understand the worst-case scenarios that could be missed

with limited funding or sampling to avoid situations such as when sampling shows

conditions are acceptable and no remedy is required, but later an acute toxicity event occurs

that is unacceptable on a population/community level. The request included the

appropriate way to design a study to predict, or hit, the catastrophic events that are

unacceptable at hazardous waste sites.

Dominic DiToro stated that for the five SEM/AVS metals (i.e., Cd, Cu, Ni, Pb, and Zn),

uptake reactions with iron sulfide (FeS) are so fast that kinetics are not an issue (as long as

time scales are greater than about 1 hour). Not much is known about oxidation processes

except that stability appears to vary by metal and tends to be generally slow. With regard

to organic matter, there appears to be two phases: a rapidly sorbed component that goes

into equilibrium within hours, and a more resistant component that requires weeks to

months to reach equilibrium. Evidence that natural, undisturbed sediments are in

equilibrium is primarily indirect; apparently there is no way to directly test this. Nothing is

known about wetting and drying systems (i.e., tidal wetlands); all information to date is

from fully submerged systems.

80
AVS Proceedings

Suggested table of factors effecting sediment equilibrium and the scales over which they may

operate.

Temporal Scale Areal Scale

Sediment

Depth (cm)

Change in

Bioavailable

Concentration

day-wk wk-mon. mon-yr cm2 m2 km2 1-5 5-25 >25 <3x 10x >10x

Input of contaminant

mass

Removal of

contaminant mass

Dilution of IW

Input of particulate

organic carbon

Mineralization of

organic carbon

( in quality)

Metabolism of organic

carbon

( in quantity)

Input of co-solvents

Anoxia in bottom

water

 pH in IW

Salinity in IW

Alkalinity in IW

Sediment stability

Inundation

Exceed solubility

The preceding discussion lead to the following potential Consensus Statement:

“It appears that in most cases undisturbed sediments can be considered as

being at equilibrium; however, there are situations and processes that may

violate the equilibrium assumption and that one must be knowledgeable

and wary of. Note that measurements cannot really be made better than

about a factor of two; equilibrium should be considered within this

81
AVS Proceedings

reality”. [While the concept behind the statement was generally agreed

upon, consensus could not be reached on the precise wording of the

statement to make it acceptable to everyone.]

Question 2: Are the following assumptions valid: (1) pore water is the sole route of

exposure to contaminants, (2) porewater exposure is equivalent to

surface water exposure?

The panel opened this discussion with the statement that pore water is not the only route of

exposure to contaminants; ingestion is also an important route for some organisms. The

route really does not matter because they both come to equilibrium. However, there are

situations where it may be important, such as for suspension-feeding clams that may never

be exposed to the sediment-bound contaminants. [Note: there was no consensus as to whether

sediment ingestion effected EqP theory.] The second issue is does it matter whether pore water

is the sole route of exposure.

Peter Landrum noted that Diaporeia ingests sediments and this is an important uptake

route, such that much higher tissue concentrations occur than would be predicted from pore

water alone. The issue becomes application of the sediment quality criteria. Do they apply

just to the infaunal deposit feeders, or to all benthic organisms regardless of different

feeding strategies (deposit feeders versus filter-feeders) or distributions within the sediment

columns (epibenthic organisms versus deep-dwelling species)? During development of the

sediment quality criteria for non-ionics, BASFs (the lipid-normalized tissue burden divided

by the carbon-normalized sediment concentration) were generated for a variety of species

and feeding strategies that clearly had different exposures. Results of the analyses

indicated that the BASFs were no different; the same mean variability in BASF was

obtained for filter-feeding sediment dwellers as for burrowing, sediment-feeding organisms.

Although the issue may still be unresolved, it was suggested that there is no compelling

reason why one should reject the idea that filter feeders are different. There was no

consensus to this idea; many participants felt that the BASF data are so highly variable that

you can’t say.

The discussion turned to the “total” exposure that includes aqueous exposures and

exposure through food sources. Dominic DiToro suggested that it doesn’t matter whether

the organism gets the dose from food or the water, because equilibrium becomes the major

issue. He questioned whether the approach is applicable to organisms of different feeding

modes and suggested the best way to answer the question is by analyzing animals of

82
AVS Proceedings

different feeding strategies. By looking at field data, we see a lot of variability, but get the

same trends regardless of feeding strategy. Dominic DiToro stated that EqP does cover

exposure through ingestion; it is not a philosophical argument, but it is based on data.

Robert Dexter asked whether different organisms placed at the same site and allowed to

come to equilibrium would have the same concentration of contaminants in their tissues.

Dominic DiToro responded that they would, with a large degree of variation associated

with the measurements.

The second part of the question of whether pore-water exposure is equivalent to surfacewater exposure was interpreted in two ways:

(1) Is exposure of an organism in sediment equivalent to that in the overlying water?

(2) Is exposure of an organism in extracted porewater equivalent to exposure of the

same organism in “spiked” dilution water.?

The workshop director clarified the question by stating that we were interested in what’s

happening in the environment, i.e., if the effects were the same for organisms that were

exposed to pore water as for organisms that were inhabiting the overlying water. Possible

differences (for both 1 and 2) pointed out by the Panel relative to the above were:

❐ More complex interactions in pore water,

❐ More ammonia (and possibly other confounding parameters) in pore water.

❐ More humic material (=OC) in pore water.

❐ More color (i.e., optical alteration) in pore water.

❐ More bacteria in pore water, less UV impact.

Possible differences relative only to (1) above are:

❐ Less diurnal DO variation in pore water.

❐ Less short-term salinity variation in pore water.

❐ Different phase partitioning processes.

❐ More refuge possibilities in pore water.

It is more probable for the organism to have refuge from a pore-water concentration that it

doesn’t like than from a surface-water concentration that it doesn’t like. However, this may

or may not be true for animals that go between the surface water and pore water.

83
AVS Proceedings

The point was made that the wording “surface water” is incorrect; the comparison is

actually with overlying water. Surface water could be mistakenly thought to be the water at

the interface with the atmosphere, which may well be more similar to pore water (because

both are interfaces).

The Panel argued that the answer to part 2 of the above question is yes if the comparison is

made based on free contaminant concentrations; no if one fails to take into account the

binding phases. There was a lot of discussion and no major disagreement, but no attempt

was made to make this a consensus statement. It was acknowledged that measuring pore

water is the weakest link in the process; it is a difficult procedure and not readily

accomplished by all laboratories.

Question 3: What effect does the presence of other organic compounds and

partitioning phases have on the bioavailability and toxicity of individual

nonionic organics?

There was general discussion on this issue. The Panel summarized that DOC does effect

bioavailability and porewater concentrations can be higher than the “free” concentrations,

but the latter are what are of concern, which is covered by the EqP theory. There was no

disagreement on the above.

84
AVS Proceedings

Panels III and IV Discussions:

Walter Berry

SAIC

Peter Landrum

NOAA

John Mahony

 Manhattan College

Burton Suedel

EA Engineering, Science & Technology

Jack Word

Battelle MSL

Question 1: Both SEM/AVS and EqP are normalization processes and require the

measurement of other components. Can these other components be

reliably and/or accurately measured?

Yes. The most difficult problems are in sample preparation as compared with actual

methodology, for example, when you have wood chips or other large materials present.

There are methods available for all necessary components, at least for sediments within a

“normal” range of organic carbon for instance. Outside this range, there is less certainty

simply because of lack of experience with these sediments.

A question arose regarding the ability of contract analytical laboratories to come close to the

standards of research laboratories who have done most of the work to date. John Mahony

believes that contract laboratories can reach and maintain the requisite standards as long as

they follow standard procedures and regularly run a standard (e.g., for AVS the standard is

a sodium sulfide solution). It is possible to make consistent and accurate AVS

measurements as long as good analytical chemical procedures are followed. Concern was

repeated regarding potential problems with contract laboratories and non-standard

sediments found at Superfund sites (e.g., sediments that contain ground coke, etc.) and

guidance was requested regarding specific QA/QC for AVS. The need for standards and

standardization was repeated.

The influence of DOC in interstitial water on TOC measurements was also discussed. The

method used for DOC measurements is important and can influence the results by a factor

of two or three. Also, the handling and storage procedures for sediments can have a

85
AVS Proceedings

significant influence on the analytical results. While oxidation of sediments will have an

affect on AVS measurements, Walter Berry’s experience has been that most sediments

within the “normal” range of AVS, can be homogenized, placed into a jar without oxygen,

and held in the dark at 4°C, with little change in the AVS content. Jim Brannon however,

emphasized that for small quantities of sediment the analytical work should be done under

nitrogen. The appropriate procedures are very dependent on the quantity of sediment you

are working with.

There was an apparent consensus that, under the extreme conditions often found at

hazardous waste sites, work needs to be done on methodologies for sample collection and

handling.

Question 2: Co-occurrence of high contaminant concentrations (as measured by

chemistry) and naturally occurring toxicants (e.g., ammonia and

hydrogen sulfide). How should they be handled with respect to

sediment evaluation and remediation?

There was common agreement that when we evaluate sediments, we are not just addressing

one factor. Sediments are a mixture and contain other contaminants and natural products.

As far as naturally occurring toxicants, ammonia and hydrogen sulfide are potential

confounding factors. Whole sediment TIEs can elucidate whether these are problems, and

procedures are now available in manuals and other documents for dealing with ammonia

(e.g., the Inland Testing Manual). Handling and storage procedures are also a factor in

sediment toxicity. The longer sediments with a high organic carbon content are stored the

more likely it is that ammonia will be produced. Collection, handling, storage, and sample

preparation (i.e., method used to obtain pore water) affects everything, not only the

chemistry but the biology as well.

There was discussion regarding the relationship between un-ionized and total ammonia. It

was pointed out that there are fewer “rules” at hazardous waste sites than in “normal”

sediments. However, toxicity is still pH dependent, with more consistency in freshwater

than in salt water (based on studies to date; there is a lot of work going on in this area).

The severity of the ammonia problem was discussed at length. The Panel attempted to

make a consensus statement that ammonia is frequently a problem in freshwater surface

sediments, and less of a problem in marine sediments unless the samples are taken from

deep sediments. There were major disagreements on this point, with some individuals

arguing it is a major factor in freshwater in shallow sediments but less so than in marine

sediments, and others stating that there is not a major problem in freshwater. The issue of

86
AVS Proceedings

sediment storage was again discussed. The 10-day, 6-week, and 8-week holding times, as

recommended in the dredge sediment evaluation guidance manuals, were questioned. The

development of a very dark, black surface layer on sediments stored at 4°C suggests

something is happening to the sediments. Whether it’s tied to ammonia or sulfides is

uncertain. However, the general feeling was that less manipulation and shorter holding

times were better.

Very little information appears to be available regarding hydrogen sulfide, which may well

not be a problem in “normal” sediments with low to moderate contamination. However,

hazardous waste sites may well be very different (though this needs to be determined).

Question 3: When multiple contaminants are present in toxic sediments can

SEM/AVS and/or EqP identify the contaminant(s) responsible for the

toxicity?

The basic answer to this question was “no”. These methods can provide insights but, by

themselves, do not provide the whole answer. The toxic components should be identified

by other methods, such as a TIE. A caution was added that the proper way to use

SEM/AVS with multiple metals, which is typical of most hazardous waste sites, is to add

all five metals (i.e., Cu, Cd, Pb, Ni, and Zn) together and compare that value to the AVS.

The individual metals should not be normalized to AVS. Iron, manganese, and Cr are not

included in the comparison. Another caution was added: when solubility limits are

approached, EqP will not be predictive.

Question 4: What are the components of a sampling plan to evaluate contaminant

bioavailability based on SEM/AVS and EqP?

The sampling plan should reflect the questions being asked. Sampling plan development

should consider the following.

(1) Exposure conditions (organisms, depths, sediment composition). The kind of

exposures to be concerned with must be identified in order to sample the

appropriate matrix, both for depth of the sediment to be collected and depth of

the sediment to be concerned about. The biology and chemistry must be matched

to the exposure conditions.

(2) Appropriate sampling equipment: the allowable degree of disturbance must be

known when deciding which type of equipment to use. The sample integrity must

reflect the initial questions being asked.

87
AVS Proceedings

(3) Preliminary characterization (pH, Eh): what can be done at the site to obtain data

that helps us understand the situation and provides some sort of baseline

information for comparison with future test results.

(4) Sample protection: what precautions during manipulation and shipment should be

taken to ensure sediment integrity? Temperature is a variable that should be

closely watched as it directly affects sediment integrity. There was a consensus

that the more the sediments are manipulated, the less it can be expected that they

resemble what is in the field.

There was detailed discussion regarding the pros and cons of compositing, field replicates

versus laboratory replicates, etc. The only conclusion was that resolution of these issues

depends on the question(s) being asked.

Question 5: What are the logistical variables that need to be identified and

standardized (e.g., sample storage, handling, acid concentration, sulfide

evolution rate, analytical considerations)?

The following factors were identified as important in the field:

❐ Sample storage: for SEM/AVS it is recommended that sediments be stored at 4°C;

sediments should not be frozen as this may effect other critical aspects of

sediment; storage should be no more than 3 weeks; generally you should test as

soon as possible.

❐ Sample handling: the appropriate container must be used, i.e., glass for AVS with

Teflon top; the sample integrity should be maintained (i.e., little to no mixing or

homogenization) with little head space in the jar; the samples should be

immediately refrigerated; if available, a large glove box with a Nitrogen atmosphere

is recommended.

Once in the laboratory, other factors become important, including:

❐ Acid type/concentration: 1M hydrochloric is recommended as the standard,

although HCl can’t always be used; some metals (i.e., silver [Ag]) precipitate in the

presence of chloride ions; nitric acid can be used in this case; phosphoric acid (but

not with Ag) is used by some. HCl is appropriate for the five AVS metals; HCl is

also appropriate for Hg2Cl but not HgCl.

❐ Sample size: 1 to 2 g is recommended for AVS/SEM analyses.

❐ Digestion time: 30 minutes is recommended for most sediments; however, if the

material in the reaction vessel has not settled or there is no distribution of color,

then additional time is required.

❐ Daily standards: the need for calibrated standards was again highly stressed;

standards should be used for all analyses (sodium sulfide for AVS analyses).

88
AVS Proceedings

The discussion did not attempt to address all the factors that should be standardized. John

Mahony pointed out that many of these issues fall into the realm of “alchemy,” that is,

experienced investigators know what they are doing and do not follow the book. However,

laboratory personnel must be effectively trained to do these analyses correctly. A lot of

discussion followed regarding problems getting contractors to accomplish this. Also, it was

agreed that there is a need for a well-defined standing operating procedure (SOP) that lists

the qualifications for contractors and provides accurate and complete guidance for these

contractors (including the use of standards). With this type of document, it could be

determined if contractor’s work was done properly.

The following variables should always be measured whenever SEM/AVS and EqP are being

used: pH, salinity, TOC, grain size, and sulfide.

Question 6: How does sampling alter sediment toxicity?

This question was not directly addressed due to lack of time, but had been part of many of

the previous discussions. With regard to AVS, it was pointed out that PRPs should be

aware that lack of care will make matters “worse” for them, not better.

89
AVS Proceedings

90
AVS Proceedings

Panel V Discussions

Peter Chapman

EVS Environmental Consultants

Dominic DiToro

HydroQual, Inc.

Spyros Pavlou

 Foster Wheeler Environmental Corporation

Question 1: How much additional field validation of the approaches is required?

Question 2: What type(s) of validation should be pursued?

Question 3: What would a field validation program consist of to validate SEM/AVS

and EqP?

The Panel suggested that they use this session to make a few initial comments regarding

these questions and then open the floor to general discussion and questions.

In response to question 1, the Panel proposed that the use of the word “field” is

inappropriate. There are a whole variety of tools; validation can take many forms and we

shouldn’t restrict it just to field validation. We need to take a look at how the tools

interact. When we look at AVS or EqP, have we done enough validation in terms of

laboratory studies to make it possible to go out in the field and do “field” validation?

What does field validation mean?

A brief presentation was made regarding the use of field transplant studies to validate

results obtained in laboratory studies. The approach is based on complementary studies

that include a combination of theory and predictions and laboratory bioassay and field

exposures, which generate a preponderance of evidence to help answer the validation

question. With this type of triad it is possible to validate, in some way at least, the

predictions until you are able to make a better case for what you’re trying to demonstrate.

There is a need for validation, but validation must be done a little bit craftily at this time. In

the past, the approach has been to do field experiments and then try to relate some of the

laboratory test results with the field. There are two primary issues: validation of the

science and validation of the methodologies to prove that the science works. The

implementation of the tools is an important issue and should be separated from the science

of the tools.

91
AVS Proceedings

There was general agreement that we should test and rationalize differences between

methods, not simply take adversarial positions. To accept a validation methodology, we

should do it in a way that we test and rationalize differences in methods, rather than

focusing on a “preferred” method. The experimental design is extremely important; we

should design the study to test the methods, because it is possible to configure an

experiment that will give the desired answer. It was proposed that we perform the tests in

a way that allows comparison of the results, with all conditions being equal. With this

approach we can determine whether or not there is consistency in the data. The issue then

becomes understanding why there is no consistency.

The issue of uncertainty was again discussed. What are the bounds? During the

discussions presented over the past few days, factors of two, five, and ten have been

proposed with respect for interpretation of data. Is an SEM/AVS ratio of 0.75 different

from 1.2? How far can one push the boundaries of the methods (e.g., wood chips in

sediments)? What works and what does not? The allowable mortality must also be

defined. Effective management decisions can only be made by comparing the variability

both within and between data sets and to do this we need to define what is allowable. A

similar concern was expressed for organic carbon. When are we pushing the uncertainty

levels? It was requested that we put some bounds on the uncertainty for the various

parameters.

The issue of spatial variability was discussed again. From a probabilistic view, it was

suggested that the number of samples collected for analysis be sufficient to justify a

probability distribution. Therefore, the data (e.g., AVS data) will have some uncertainty

based on the way the sample was collected. The probability approach must be used

carefully; the goal is to collect enough samples that permit you to make some sense of the

data.

The bioaccumulation issue should be discussed in greater detail (perhaps a following

workshop). This should include the spatial averaging of exposure. Many cleanup and

implementation decisions are based on the upper-level food consumption and risk, and we

must accurately define the risk to exposed biota. To accomplish this, it was suggested that

the screening tools (i.e., AVS and EqP) be linked to the risk assessment process and,

ultimately, to risk management. Spyros Pavlou indicated that the Marine Board, of which

he is a member, is in the process of incorporating contamination assessment with risk

management decisions to implement an action and develop the most cost-effective actions.

92
AVS Proceedings

He further suggested that the members of the workshop link the risk assessment and the risk

management processes with the assessment methods.

Dominic DiToro noted that it is extraordinarily difficult to actually do a field validation; in

fact one can only do a field invalidation. We don’t know why it worked, if it worked; we

only have concurrence. Exactly how to validate is the major question. We haven’t done a

good job with validation because we don’t know how to validate very much. The ultimate

test is the arena of utility. We have utility if more and more investigators find that the

methods are useful in interpretation of their data. If the methods fade because of conflicting

or uninterpretable results, then the methods will be forgotten. However, there is one

observation that can be made that will invalidate the AVS/EqP methodology, and that is if

there is a co-occurrence of an organism and a concentration that is predicted to kill that

organism. That becomes an invalidation. The absence of an organism might be for any

reason, but the co-occurrence strongly indicates invalidation.

The problems and questions that we are faced with are both confusing and frustrating

because nobody knows all the answers, and we don’t know how to find the answers. How

do we do an ecological risk assessment? If we look in the tool box, we find that it’s

embarrassingly bare. It is undeniably true that there is no one way to answer these

questions “yes” or “no,” “pass” or “fail”. There is no single tool that will give all the

answers (this includes AVS and EqP). Compared to other well-developed sciences, the

science of environmental risk assessment is in its infancy. When something does work, it

then becomes the topic of microscopic examination (i.e., SEM/AVS). However, the general

feeling was that at least the community was interested in the approach, even though there

may be some disagreement as to its utility.

There was an extensive discussion that mixed policy and science (e.g., sediment quality

criteria based on AVS and EqP). Dominic DiToro described the uncertainty analysis

conducted on the EPA criteria based on the EqP prediction methodology. The analysis was

conducted on all the data sets used to develop the criteria. The analysis tested the ability

to predict sediment toxicity tests from independent measurements (i.e., water only

measurements and independent estimates of octanol water partition coefficients). The EPA

criteria development process included field validation efforts. Rick Swartz pointed out that

one should go to the field earlier rather than later, and that some field sites will be easier to

work with than others (i.e., field validate at the simplest sites first). One of the most

powerful tools for field validation is colonization experiments (e.g., placing trays of

sediment back in the environment). A suggestion was made for conducting work with

93
AVS Proceedings

undisturbed cores; dosing can be done in micro- or mesocosms. Peter Chapman pointed out

that validation can be correlative (e.g., observations of the benthos related to other “tools”)

or investigative (e.g., undisturbed cores, trays of contaminated sediment put back in the

field, etc.). He also recommended that, given that hazardous waste sites are only consistent

in their inconsistency, there be an attempt to piggyback useful validation-type studies in

hazardous waste site investigations. But these will only be useful if they are reported.

Dominic DiToro clarified the application of EqP. There is a misunderstanding about the

range of its applicability, with some investigators applying the approach to large organisms

like fish. They then make the assumption that the fish is in equilibrium with the sediment.

The EqP theory only works for benthic organisms, and even for that specific group of

organisms there is a lot of scatter in the data. However, it has general applicability.

Spyros Pavlou reiterated that the SEM/AVS process is a screening tool that may represent

the worst-case scenario. There are many more things that go into evaluation of Superfund

sites than AVS or EqP. They should be used to identify the areas that require further

investigation. The uncertainty associated with either technique becomes less important as

more tools are brought into the evaluation process. However, if the decision is based solely

on one approach, then you must know the uncertainty associated with that approach.

There was a request that both biology and chemistry be used to make the decisions; neither

should be used alone. AVS and EqP should be used for screening, to help set priorities, and

to assist in determining cause-and-effect.

94
AVS Proceedings

NOAA’s Overview and Summation

The Woods Hole Oceanographic Institute was a stimulating venue for the open exchange of

information. The many theories and viewpoints presented over the course of the workshop

lead to several open discussions that gave all participants an opportunity to express their

concerns and opinions. The workshop was successful in achieving the goals as set forth by

NOAA. Clarification was provided regarding:

(1) the scientific theory on which EqP and AVS methodologies are based,

(2) the utility of EqP and AVS in the evaluation of contaminated sediments, and

(3) the application of these approaches at hazardous waste sites.

There was overwhelming agreement that the EqP and AVS methodologies have a place in

the risk assessor’s tool box, but these tools must be properly used with the investigator

understanding and implementing all associated caveats.

The presenters discussed both the strengths and weaknesses of the methodologies as well as

our ability to accurately characterize aquatic systems. Overall, there was general agreement

on many aspects of the methodologies, but there were many areas where disagreement

prevailed suggesting the need for additional research. During the early stages of the

workshop it was suggested that we attempt to reach consensus on as many AVS/EqP

issues as possible. Candidate consensus items were recorded and edited as new

information became available. Consensus was reached when the vast majority, if not all, of

the 24 participating scientists agreed. The following six consensus items resulted from the

AVS Workshop.

Consensus Item #1:

Bulk chemistry determinations alone are generally relatively poor

predictors of contaminant bioavailability or toxicity.

As with many other tools in the investigator’s toolbox, the utility of bulk sediment analysis

is limited when evaluating hazardous waste sites. Bulk sediment analysis data can be used

to identify the presence and distribution of contaminants within an aquatic system and

suggest which ones may be of major concern. Also, because of management and regulatory

issues effecting hazardous waste sites, bulk sediment data still may ultimately be the

measure by which areas requiring remediation are delineated. However, these measurements

are not predictive of bioavailability, toxicity, or potential threat to aquatic biota. The

95
AVS Proceedings

realization that bulk chemistry determinations could not provide information regarding

contaminant bioavailability or toxicity has prompted researchers to investigate other

approaches. The EqP and AVS methodologies are two attempts to provide chemistry data

that can either help predict or explain sediment toxicity. It must be remembered that these

methodologies do not replace bulk sediment analysis at hazardous waste sites, but rather

supplement it. This is because remediation decisions are not just based on whether the

sediments are toxic in situ, but also on whether they may act as a source for downstream

contamination.

Consensus Item #2:

Pore water is the primary route of exposure to many contaminants for

benthic organisms; however, ingestion is also an important route for

some organisms.

The only methods by which benthic organism can uptake contaminants are active or passive

absorption through the gills or skin (including the gut lining). For absorption to occur, the

contaminant must be in a dissolved form either in the surrounding pore water or in the gut.

Contaminants enter the gut via the ingestion of pore water or sediments; contaminants

associated with sediments may become available for absorption as a result of pH

conditions and metabolic processes. While no consensus was reached on whether sediment

ingestion effects the ability of EqP and AVS theory to predict bioavailability and/or

toxicity, Dr. DiToro expressed the view that the metal sulfide bonds are too strong to be

broken by any activity in an organism’s gut so sediment ingestion should have no effect on

the application of AVS theory. It is still a question with regard to the EqP theory for nonionic organic contaminants. However, the work of Meador et al. (in press) suggests that

sediment ingestion can increase the bioavailability of at least some non-ionic organic

contaminants. They compared the bioaccumulation of PAHs by two different benthic

organisms: Rhepoxynius abronius (a non-deposit feeding amphipod) and Armandia brevis (a

non-selective deposit feeding polychaete). Over a wide range of PAH exposure, no

significant difference was found between species in the tissue concentrations of low

molecular weight PAHs. There was a large differential observed between species for the

bioaccumulation of high molecular weight PAHs, with the polychaete accumulating more

than the amphipod.

96
AVS Proceedings

Consensus Item #3:

The EqP and AVS approaches to sediment analysis are based on sound

science and theory; partitioning does occur, and sulfides do complex with

some metals (Cd, Cu, Pb, Ni, Zn), affecting their bioavailability.

The third consensus item is based on the EqP theory, which states: there is a calculable

relationship between sediment contaminant concentrations and contaminant toxicity.

Therefore, if you know the water-only LC50 (or some other water-only toxic benchmark) you

can calculate a sediment LC50 for the contaminant based on its Koc and the organic carbon

content of the sediment. Data were presented at the workshop supporting this for single

non-ionic organic contaminants under laboratory conditions. (Work has yet to be done on

mixtures of non-ionic organic contaminants, the more common case in the real world.) The

theoretical basis for the AVS approach is that the sulfides of the five metals have lower

solubility constants than either Fe or Mn sulfides, and if they are added to a system

containing FeS and/or MnS, they will displace the Fe and/or Mn forming an insoluble metal

sulfide and thus make the contaminant metal non-bioavailable. Like the first, Consensus

Item #3 was readily agreed to, although there were many questions regarding how this could

be applied to the real world.

Consensus Item #4:

For sediment in which the contaminant(s) of concern include one or more

of the following: Cd, Cu, Ni, Pb, and Zn, an SEM/AVS molar ratio of <1

indicates the vast majority of these metals are complexed with sulfides

and are essentially not bioavailable and hence acute toxicity from these

metals is highly unlikely. An SEM/AVS molar ratio >1 does not

necessarily indicate that these metals are bioavailable and acutely toxic;

other factors may be affecting their bioavailability and toxicity. Toxicity

refers to laboratory tests only and does not necessarily apply to

bioaccumulation of the metals or alterations in the benthos. [Noted

Caveats: (1) The uncertainty of the measurements must be taken into

account when comparing the ratio to the absolute value of 1. (2) Both the

numerator and denominator must be considered when evaluating the

ratio: almost “no” AVS compared to “some” SEM has a different meaning

than the same ratio wherein the AVS concentration is much higher. Very

low concentrations of both SEM and AVS should not be considered

predictive. (3) The consensus statement is only valid if AVS

concentrations are above operational detection limits.]

Consensus was reached on this item concerning what the AVS methodology can and cannot

currently tell us, although it was necessary to include several caveats. This consensus item

97
AVS Proceedings

includes several important issues, each of which should be thoroughly discussed. The first

major issue is that currently there are only five metals applicable to the AVS methodology:

Cd, Cu, Ni, Pb, and Zn. Other metals are being investigated, but now there is insufficient

evidence or data to suggest they behave in a similar manner. Second, the SEM portion of the

SEM/AVS ratio is the sum of the molar concentrations for these five metals; the ratio is not

meant to be calculated on an individual metal basis. The AVS methodology is not

considered useful in evaluating Hg contamination because although inorganic Hg also forms

an insoluble sulfide (HgS), the primary concern regarding Hg is the biomagnification of

organic Hg. However, HgS is less soluble than any of the other five metal sulfides, and if

significant quantities of inorganic Hg are present they could compromise the evaluation of

the bioavailability of the other five metals by making the sulfide unavailable for binding.

The interpretation of the SEM/AVS ratio is a very important step in the sediment

evaluation process. The purpose of the ratio is to determine how much of the metals

present are bound and biologically unavailable. If the ratio is <1, theory indicates that

virtually all five metals are bound up as sulfides and are unavailable to biological receptors.

Hence, the unavailable metals are not toxic. This theory was supported by data from

toxicity tests conducted with both spiked- and field-collected sediments. A ratio >1 only

indicates that all metals are not bound up as sulfides. There are several reasons why acute

toxicity may not be observed at ratios greater than one: the metals may be bound to other

ligands and not be bioavailable; they may be in ionic form but at sufficiently low

concentrations as to not be acutely toxic, or the test organisms may actively avoid the

contaminants. The ratio by itself gives no indication as to the quantity of metals that are

bioavailable, and this is where the second caveat is of concern. For example, if the SEM

concentration is 0.6 μmoles and the AVS concentration is 0.3 μmoles, the SEM/AVS ratio is

2 with 0.3 μmoles of metals potentially bioavailable. However, if SEM=6 μmoles and

AVS=3 μmoles you still have a ratio of 2 but now you have 3 μmoles of metals potentially

bioavailable. So, the ratio is fine for indicating no toxicity, but can't tell when to expect

toxicity. At this point, some other mechanism must be used to determine bioavailability

and toxicity (i.e., field or laboratory toxicity tests).

It has been recommended that in the future the difference between SEM and AVS (SEMAVS) be used instead of the ratio (SEM/AVS). The difference is more informative than the

ratio. As in the above hypothetical example, the difference would provide information on

the amount of bioavailable metals. Also, if the difference is calculated sequentially for the

five metals, starting with the one with the least soluble sulfide, the metal(s) that is actually

bioavailable could be determined. Dr. Hansen expressed the opinion that, based on his

98
AVS Proceedings

spiked- sediment colonization tests, in the future we will be able to predict chronic effects

through the use of SEM-AVS in conjunction with our understanding of the role of organic

carbon.

Consensus Item #5:

The quantity and type of organic carbon can affect the bioavailability of

non-polar organic contaminants as well as some metals.

It was readily agreed that the quality and quantity of organic carbon found in the sediments

affects the bioavailability of non-polar organic contaminants. Dr. Mahony indicated that

once the AVS are used up via complexation, the organic carbon can strongly affect the

bioavailability of Cu, and to a lesser extent Pb. With regard to the importance of the type

of organic carbon present, it was generally agreed that humic materials were common in

most sediments and subtle variations in the quality would probably have little effect on

contaminant availability and Koc. However, contaminant availability could be influenced

by the presence of large masses of organic carbon, such as coal, wood chips, creosote, or

dense plant material. The presence of foreign material and large chunks of debris must be

kept in mind when analyzing and interpreting chemical data for sediments. It was also

noted that hazardous waste sites are special cases where the type of organic carbon may be

an important factor in determining bioavailability of contaminants.

Consensus Item #6:

There is no single, 'perfect' tool that should be used in isolation for

making decisions. This includes, for example, single-chemical criteria,

toxicity tests, and faunal studies. The best decisions are based on the

collection, use, and fundamental understanding of biological and chemical

data.

This last consensus item is possibly the most important from a risk assessor's or manager's

viewpoint. It provides the limits and boundaries of these methodologies and enforces the

concept of a multi-faceted approach. This consensus item puts a perspective on the overall

risk assessment process and forces us to realize that EqP and AVS are not “magic bullets”

that will provide the ultimate answer. However, they can be useful in the evaluation

process for contaminated sediments at hazardous waste sites, if they are used properly

with full knowledge of their limitations and in conjunction with the other available tools.

A seventh consensus item could be listed, although it was not formally presented as such at

the workshop it was repeatedly emphasized:

99
AVS Proceedings

“always be aware of the uncertainty inherent in your analyses.”

Agreement within a factor of two is considered very good for any replicate environmental

measurements, whether it be chemical or biological. Some consider a factor of five or even

an order of magnitude as not unreasonable depending on the parameter measured. This

uncertainty is not only due to our limited ability to make certain types of measurements, but

is also due to the natural variability found in the real world. When analyzing any data set

or the output of any model, decision makers must be aware of the uncertainty associated

with the values they are attempting to evaluate. Awareness of uncertainty can be especially

important if sample data are being compared to any form of criteria. If the criterion is an

SEM/AVS ratio of 1, is 0.9 <1 and is 1.9 truly >1? Only if the decision makers are aware

of the uncertainty can they be expected to draw reasonably appropriate conclusions about

conditions at a site.

Although the group was able to reach consensus on six issues, a number of disputed and

unresolved issues still remain, including:

 “Does equilibrium exist in natural sediments? How does EqP work with

mixtures? How does sediment ingestion effect the EqP analysis? What are

the proper procedures for collecting and preparing sediment samples for

SEM and AVS analyses?”

A consensus statement was almost reached regarding equilibrium in natural sediments. The

predominant belief was that within our ability to measure, equilibrium (or at least steadystate conditions) exists in natural, undisturbed sediments. In this case, equilibrium is

defined for the contaminant in the appropriate solid phase of the sediment and the pore

water, not for the solid phase or pore water and the biota. However, due to lack of

resolution on what constitutes undisturbed sediments, the recognition that hazardous waste

site sediments are highly disturbed and may receive continuing inputs, the special situation

of periodically inundated sediments, and several other questions and caveats, a consensus

could not be reached.

No data were presented to address either the mixtures or ingestion questions. As for

sediment sample handling, it was suggested that if done with reasonable care; samples

could be homogenized with only a minimum lose of AVS because the contaminant metal

sulfides are less easily oxidized than the Fe or Mn sulfides. It was also stated that if

relatively small samples were being collected, then any handling should be done under a

100
AVS Proceedings

nitrogen atmosphere. There was general agreement on the need for specific protocols for

sample collection, handling, and analysis.

Another question asked was:

 “Since AVS is only present under anoxic conditions, how does it effect the

oxic sediments where the biota live?”

The basic theory was presented that the free metal ions in the pore water will migrate from

the region of relatively high concentration (oxic sediment pore water) to the region of

relatively low concentration (anoxic sediment pore water) where they will be removed

through the formation of metal sulfides. This was supported by incremental porewater

concentration curves for collected sediments in the laboratory and in situ sediments. Both

sediments showed similar conditions. A possible way around this question is the work Dr.

Tessier presented on determining a calculable relationship between metal (Cd)

concentrations in oxic sediments and in biota. It would be interesting to see if there is a

relationship between what he is doing in oxic sediments with what is being done with AVS

in anoxic, i.e., are the oxides he is measuring a direct indicator of the sulfides that would be

present in the anoxic layer.

In conclusion, we find that the EqP and AVS approaches to sediment analysis are among

the tools that can be used to evaluate sediment contamination at hazardous waste sites.

However, they only should be used with a full understanding of their advantages and

limitations, and before they can be used to their fullest potential several questions need to

be answered. Finally, as was reiterated throughout the workshop, there is no miracle tool or

single number that gives all the answers; AVS and EqP, like any of the other available tools

(e.g., bulk chemistry, toxicity tests, community studies), should not be used in isolation.

Decisions at a hazardous waste site need to be based on an integrated approach using both

site-specific biological and chemical data.

101

AVS Proceedings

Appendix A

Workshop Participants

Participants

Name Affiliation

Jack W. Anderson Columbia Analytical Services

Carlsbad, California

Walter J Berry Science Applications International Corp.

Narragansett, Rhode Island

James M. Brannon USACOE Waterways Experimental Station

Vicksburg, Mississippi

R. Scott Carr National Biological Survey

Corpus Christi, Texas

Peter M. Chapman EVS Environmental Consultants

North Vancouver, British Columbia

Theodore H. DeWitt Battelle Marine Science Laboratory

Sequim, Washington

Robert N. Dexter EVS Environmental Consultants

Seattle, Washington

Dominic M. DiToro HydroQual, Inc.

Mahwah, New Jersey

David J. Hansen US EPA Environmental Research Lab

Narragansett, Rhode Island

Christopher G. Ingersoll National Biological Survey

Columbus, Missouri

D. Michael Johns EVS Environmental Consultants

Seattle, Washington

James L. Lake US EPA Environmental Research Lab

Narragansett, Rhode Island

Peter F. Landrum NOAA Great Lakes Environmental

Research Laboratory

Ann Arbor, Michigan

A-1
AVS Proceedings

Name Affiliation

Brenda K. Lasorsa Battelle Marine Science Laboratory

Sequim, Washington

John D Mahony Manhattan College

Riverdale, New York

James P. Meador NOAA National Marine Fisheries Service

Seattle, Washington

Spyros P. Pavlou Foster Wheeler Environmental Corp.

Stuart, Florida

Philippe Ross The Citadel

Charleston, South Carolina

K. John Scott Science Applications International Corp.

Narragansett, Rhode Island

Mark D. Sprenger US EPA Environmental Response Team

Edison, New Jersey

Burton Suedel EA Engineering, Science & Technology

Hunt Valley, Maryland

Richard C. Swartz US EPA Environmental Research Lab

Newport, Oregon

André Tessier University of Quebec

Sainte-Foy, Quebec

Jack Q. Word Battelle Marine Science Laboratory

Sequim, Washington

A-2
AVS Proceedings

NOAA Coastal Resource Coordination Branch Staff

Alyce Fritz – Branch Chief Waynon Johnson

Gail Beckman Peter Knight

Trey Brown John Lindsay

Mike Buchman Don MacDonald

Chuck Cairncross Mary Matta

Herb Curl Chris Mebane

Jay Field Greg Morris

Ken Finkelstein Natalie Sanbe

Ron Gouguet Mike Wenholz

Other Attendees

Name Affiliation

Dave Charters US EPA Environmental Response Team

Edison, New Jersey

Chris Hall Genwest

Seattle, Washington

Jackie McGee Genwest

Seattle, Washington

Nancy Morse US Fish and Wildlife Service

Edison, New Jersey

Royal Nadeau US EPA Environmental Response Team

Edison, New Jersey

Mary Reiley US EPA Office of Science & Technology

Washington, D.C.

Sandra Salazar EVS Environmental Consultants

Seattle, Washington

Susan Svirsky US EPA Region I

Boston, Massachusetts

Dave Tomey US EPA Region I

Boston, Massachusetts

A-3

AVS Proceedings

Appendix B

Agenda

Tuesday September 13

Moderator: Donald MacDonald

 9:00 - 9:50 Greetings and opening comments Alyce Fritz

Donald MacDonald

 9:50 - 10:20 Assessing the assessment tools Robert Dexter

EVS Environmental

Consultants

10:20 - 10:35 Break

10:35 - 11:10 An introduction to EqP applied to nonionic organic

chemicals and metals

Dominic DiToro

HydroQual, Inc.

11:10 - 11:45 Generic screening versus site-specific reality: What

works, what doesn't and why

Peter Chapman

EVS Environmental

Consultants

11:45 - 1:00 Lunch

Moderator: Mary Matta

 1:00 - 1:30 Certain chemical aspects of heavy metal-sediment

interactions: the role of organic carbon and the kinetics of

metal sulfide oxidation

John Mahony

Manhattan college

 1:30 - 2:00 The influence of sediment stability/disturbance on

toxicity of non-polar organic compounds to marine

amphipods

Jack Word

Battelle MSL

 2:00 - 2:30 KOC and KDOC in sediment pore water Jim Brannon

US Army COE WES

 2:30 - 3:00 Pore water toxicity testing the direct application of EqP Scott Carr

US NBS

 3:00 - 3:15 Break

Moderator: Alyce Fritz

 3:15 - 3:45 Prediction of Cd accumulation in benthic organisms André Tessier

University of Quebec

 3:45 - 4:15 Predicting the toxicity of metals-contaminated sediments

using AVS and interstitial water normalizations: results

from spiked sediment laboratory ten-day lethality and

four-month colonization tests

David Hansen

US EPA ERL

 4:15 - 4:45 "Marathon Battery: What were the questions?" Mark Sprenger

US EPA ERT

 4:45 - 5:15 Sediment contamination, toxicity and macrobenthos:

comparability of laboratory and field observations

Rick Swartz

US EPA ERL

B-1
AVS Proceedings

Wednesday September 14

Moderator: Jay Field

 8:00 - 8:30 Use of SEM/AVS ratios and pore water metal

concentrations to predict metal bioavailability in field

sediments

Gerald Ankley

US EPA

 8:30 - 9:00 Comparison between the results of toxicity testing with

marine sediments and predicted effects based on chemistry

Jack Anderson

Columbia Analytical

Services

 9:00 - 9:30 Assessing toxicity in multi-metal contaminated sediment:

Application of theoretical and investigative approaches

Michael Johns

EVS Environmental

Consultants

 9:30 - 10:00 Strengths and weaknesses of the EqP approach as applied

to hazardous waste sites risk assessment

Spyros Pavlou

Foster Wheeler

Environmental Corp.

10:00 - 10:30 Break

10:30 - 12:15 Panel Discussion I

Moderator: Mary Matta

Panelists

1. What can and cannot the SEM/AVS ratio really tell

us?

2. How do the temporal (e.g., seasonal) and spatial

changes in physico-chemical parameters affect the

SEM/AVS ratio and how does this variability affect our

assessment of hazardous waste site conditions?

3. Which parameters of the system should be measured

and how can this data be applied to an entire ecosystem?

Chris Ingersoll

NBS

Brenda Lasorsa

Battelle MSL

John Scott

SAIC

12:15 - 1:15 Lunch

1:15 - 3:00 Panel Discussion II

Moderator: John Lindsay

Panelists

1. Is equilibrium a good assumption in the environment?

2. Are the following assumptions valid?

a. pore water is the sole route of exposure to contaminants;

b. pore water exposure is equivalent to surface water

exposure.

3. What effect does the presence of other organic

compounds and partitioning phases have on the

bioavailability and toxicity of individual nonionic

organics?

Theodore DeWitt

Battelle MSL

Jim Meador

NOAA, NMFS

Philippe Ross

The Citadel

 3:00 - 3:15 Break

B-2
AVS Proceedings

Wednesday September 14

 3:15 - 5:00 Panel Discussion III

Moderator: Mike Buchman

Panelists

1. Both SEM/AVS and EqP are normalization processes

and require the measurement of other components. Can

these other components be reliably and/or accurately

measured?

2. Co-occurrence of high contaminant concentrations (as

measured by chemistry) and naturally occurring toxicants

(e.g., ammonia and hydrogen sulfide) -- how should they

be handled with respect to sediment

evaluation/remediation?

3. When multiple contaminants are present in toxic

sediments can SEM/AVS and/or EqP identify the

contaminants responsible for the toxicity?

Walter Berry

SAIC

John Mahony

Manhattan College

Jack Word

Battelle MSL

Thursday September 15

 8:00 - 9:45 Panel Discussion IV

Moderator: Alyce Fritz

Panelists

1. What are the components of a sampling plan to evaluate

contaminant bioavailability based on SEM/AVS and EqP?

2. What are the logistical variables that need to be

identified and standardized (e.g., sample storage,

handling, acid concentration, sulfide evolution rate,

analytical considerations)?

3. How does sampling alter sediment toxicity?

Peter Landrum

NOAA Great Lakes ERL

Burton Suedel

EA Engineering, Science &

Technology

 9:45 - 10:00 Break

10:00 - 11:45 Panel Discussion V

Moderator: Don MacDonald

Panelists

1. How much additional field validation of the approaches

is required?

2. What type(s) of validation should be pursued?

3. What would a field monitoring program consist of to

validate SEM/AVS and EqP?

Peter Chapman

EVS Environmental

Consultants

Dominic DiToro

HydroQual, Inc.

Spyros Pavlou

Foster Wheeler

Environmental Corp.

11:45 - 12:15 Wrap-up

B-3
AVS Proceedings

B-4
AVS Proceedings

Appendix C

References

The following list contains references dealing with AVS and EqP as well as other topics

related to the workshop. It includes all references cited by the various presenters. The list is

not meant to be comprehensive; it is a starting point for anyone wishing to explore the

subject in more detail than was covered by this summary of the proceedings.

Adams, W.J., R.A. Kimerle, and J.W. Barnett, Jr. 1992. Sediment quality and aquatic life

assessment. Environmental Science and Technology 26: 1864-1875.

Adams, W.J., R.A. Kimerle, and R.G. Mosher. 1985. Aquatic safety assessment of

chemicals sorbed to sediments. in Seventh Annual Symposium, on Aquatic Toxicology and

Hazard Assessment. Sponsored by ASTM Committee E-47 on Biological Effects and

Environmental Fates. Milwaukee, Wisconsin, 17-19 April, 1983, ASTM STP 854. eds. R.D.

Cardwell, R. Purdy, and R.C. Bahner, Philadelphia: American Society for Testing and

Materials. pp 429-453.

Allen, H.E., G. Fu, and B. Deng. 1993. Analysis of acid-volatile sulfide (AVS) and

simultaneously extracted metals (SEM) for the estimation of potential toxicity in aquatic

sediments. Environmental Toxicity and Chemistry 12: 1441-1454.

Allen, H.E., P. van Beelen, and J. Struijs. 1993. Acid volatile sulfide (Workshop). Science

of the Total Environment, Supplement. pp 1789-1791.

Ankley, G.T., D.A. Benoit, J.C. Balogh, T.B. Reynoldson, K.E. Day, and R.A. Hoke. 1994.

Evaluation of potential confounding factors in sediment toxicity tests with three freshwater

benthic invertebrates. Environmental Toxicity and Chemistry 13: pp 627-63.

Ankley, G.T., D.J. Call, J.S. Cox, M.D. Kahl, R.A. Hoke and P.A. Kosian. 1994. Organic

carbon partitioning as a basis for predicting the toxicity of chlorpyrifos in sediments.

Environmental Toxicity and Chemistry 13: 621-626.

Ankley, G.T., G.L. Phipps, E.N. Leonard, D.A. Benoit, V.R. Mattson, P.A. Kosian, A.M.

Cotter, J.R. Dierkes, D.J. Hansen, and J.D. Mahony. 1991. Acid-volatile sulfide as a factor

mediating Cd and Ni bioavailability in contaminated sediments. Environmental Toxicity and

Chemistry 10: 1299-1307.

Ankley, G.T., M.K. Schubauer-Berigan and R.A. Hoke. 1992. Use of toxicity identification

evaluation techniques to identify dredged material disposal options: A proposed approach.

Environmental Management 16:1-6.

Ankley, G.T., P.M. Cook, A.R. Carlson, D.J. Call, J.A. Swenson, H.F. Corcoran, and R.A.

Hoke. 1992. Bioaccumulation of PCBs from sediments by oligochaetes and fishes:

comparison of laboratory and field studies. Canadian Journal of Fisheries and Aquatic Sciences

49: 2080-2085.

C-1
AVS Proceedings

Ankley, G.T., V.R. Mattson, E.N. Leonard, C.W. West, and J.L. Bennett. 1993. Predicting

the acute toxicity of Cu in freshwater sediments: evaluation of the role of acid-volatile

sulfide. Environmental Toxicity and Chemistry 12: 315-320.

ASTM. 1992. Standard guide for conducting 10-day static sediment toxicity tests with

marine and estuarine amphipods. E 1367-90. 1992 Annual Book of ASTM Standards,

Volume 11.04. Philadelphia: ASTM. pp 1083-1106.

Bergen, B.J., W.G. Nelson, and R.J. Pruell. 1993. Partitioning of polychlorinated biphenyl

congeners in the seawater of New Bedford Harbor, Massachusetts. Environmental Science

and Technology 27: 938-942.

Boulegue, J., C.J. Lord III and T.M. Church. 1982. Sulfur speciation and associated trace

metals (Fe, Cu) in the pore waters of Great Marsh, Delaware. Geochimica et Cosmochimica

46: 453-464.

Brannon, J.M., J.C. Pennington, E.M. Davis, and C. Hayes. 1995. Fluoranthene KDOC in

sediment pore waters. Chemosphere 30: 419-428.

Brannon, J.M., C.B. Price, F.J. Reilly Jr., J.C. Pennington, and V.A. McFarland. 1993. Effects

of sediment organic carbon on distribution of radio-labeled fluoranthene and PCBs among

sediment, interstitial water, and biota. Bulletin of Environmental Contamination and

Toxicology 51: 873-880.

Brouwer, H. A. and T.P. Murphy. 1994. Diffusion method for the determination of acidvolatile sulfides (AVS) in sediment. Environmental Toxicity and Chemistry 13: 1273-1275.

Bryan, G.W. and W.J. Langston. 1992. Bioavailability, accumulation and effects of heavy

metals in sediments with special reference to United Kingdom estuaries: a review.

Environmental Pollution 76: 89-131.

Burmaster, D.E., C.A. Menzie, J.S. Freshman, J.A. Burris, N.I. Maxwell, and S.R. Drew.

1991. Assessment of methods for estimating aquatic hazards at Superfund-type sites: a

cautionary tale. Environmental Toxicity and Chemistry 10: 827-842.

Burton, G.A., Jr. 1991. Assessing the toxicity of freshwater sediments. Environmental

Toxicity and Chemistry 10: 1585-1627.

Burton, G.A., Jr. 1992. Assessing contaminated sediments. Environmental Science and

Technology 26: 1862-1863.

Carlson, A.R., G.L. Phipps, V.R. Mattson, P.A. Kosian, and A.M. Cotter. 1991. The role of

acid-volatile sulfide in determining cadmium bioavailability and toxicity in freshwater.

Environmental Toxicity and Chemistry 10: 1309-1319.

Carr, R.S. and D.C. Chapman. 1992. Comparisons of solid-phase and pore-water

approaches for assessing the quality of marine and estuarine sediments. Chemistry and

Ecology 7: 19-30.

C-2
AVS Proceedings

Casas, A.M. and E.A. Crecelius. 1994. Relationship between AVS and the toxicity of zinc,

lead and copper in marine sediments. Environmental Toxicology and Chemistry 13: 529-536.

Chapman, P. M. 1989. Current approaches to developing sediment quality criteria.

Environmental Toxicology and Chemistry 8: 589-599.

Chapman, P.M. 1991. Environmental quality criteria: What type should we be

developing?. Environmental Science and Technology 25: 1352-1359.

Chin, Y. and P.M. Gschwend. 1992. Partitioning of polycyclic aromatic hydrocarbons to

marine porewater organic colloids. Environmental Science and Technology 26: 1621-1626.

Cornwell, J.C., and J.W. Morse,. 1987. The characterization of iron sulfide minerals in

anoxic marine sediments. Marine Chemistry 22: 193-206.

Davies-Colley, R.J., P.O. Nelson and K.J. Williamson. 1985. Sulfide control of Cd and Cu

concentrations in anaerobic estuarine sediments. Marine Chemistry 16: 173-186.

Davis, W.M. 1993. Influence of humic substance structure and composition on interactions

with hydrophobic organic compounds. PhD. Gainesville, Florida: University of Florida.

Dewitt, T.H., R.J. Ozretich, R.C. Swartz, J.O. Lamberson, D.W. Schults, G.R. Ditsworth,

J.K.P. Jones, L. Hoselton, and L.M. Smithe. 1992. The influence of organic matter quality on

the toxicity and partitioning of sediment-associated fluoranthene. Environmental Toxicology

and Chemistry 11: 197-208.

DiToro, D.M., C.S. Zarba, D.J. Hansen, W.J. Berry, R.C. Swartz, C.E. Cowan, S.P. Pavlou,

H.E. Allen, N.A. Thomas and P.R. Paquin. 1991. Technical basis for establishing sediment

quality criteria for nonionic organic chemicals using EqP. Environmental Toxicology and

Chemistry 10: 1541-1583.

DiToro, D.M., J.D. Mahony, D.J. Hansen, K.J. Scott, M.B. Hicks, S.M. Mayr, and M.S.

Redmond. 1990. Toxicity of Cd in sediments: the role of AVS. Environmental Toxicity and

Chemistry, 9: 1487-1502.

DiToro, D.M., J.D. Mahony, D.J. Hansen, K.J. Scott, A.R. Carlson. and G.T. Ankley. 1992.

Acid volatile sulfide predicts the acute toxicity of Cd and Ni in sediments. Environmental

Science and Technology 26: 96-101.

Emerson, S., L. Jacobs and B. Tebo. 1981. The behavior of trace metals in marine anoxic

waters: solubilities at the oxygen-hydrogen sulfide interface. in Trace Metals in Sea Water,

eds. C.S Wong, E. Boyle, K.W. Bruland, J.D. Burton, and E.D. Goldberg. New York: Plenum

Press. pp 579-608.

Ferraro, S.P., H. Lee II, R.J. Ozretich and D.T. Specht. 1990. Predicting bioaccumulation

potential: A test of a fugacity-based model. Archives Environmental Contamination and

Toxicology 19: 386-394.

Fu, G. and H.E. Allen. 1992. Cd adsorption by oxic sediments. Water Research 26: 225-233.

Fu, G., H.E. Allen and Y. Gao. 1992. The importance of humic acids to proton and Cd

binding in sediments. Environmental Toxicity and Chemistry 11: 1363-1372.

C-3
AVS Proceedings

Garbarini, D.R. and L.W. Lion. 1986. Influence of the nature of soil organics on the

sorption of toluene and trichloroethylene. Environmental Science and Technology 20: 1263-

1269.

Gardner, M. and A. Gunn. 1989. The effect of natural ligands on trace metal partitioning.

Chemosphere 19: 1251-1260.

Gauthier, T.D., W.R. Seitz, and C.L. Grant. 1987. Effects of Structural and compositional

variations of dissolved humic materials on pyrene Koc values. Environmental Science and

Technology 21: 243-248.

Grathwohl, P. 1990. Influence of Organic Matter from Soils and Sediments from Various

Origins on the Sorption of Some Chlorinated Aliphatic Hydrocarbons: Implications on Koc

Correlations. Environmental Science and Technology 24: 1687-1693.

Green, A.S., G.T. Chandler, and E.R. Blood. 1993. Aqueous-, pore-water-, and sedimentphase Cd: toxicity relationships for a meiobenthic copepod. Environmental Toxicity and

Chemistry 12: 1497-1506.

Hare, L., R. Carignan and M.A. Huerta-Diaz. 1994. A field study of metal toxicity and

accumulation by benthic invertebrates; implications for the acid-volatile sulfide (AVS)

model. Limnology and Oceanography 39: 1653-1668.

Harkey, G.A., P.F. Landrum, and S.J. Klaine. 1994. Comparison of whole-sediment,

elutriate and pore-water exposures for use in assessing sediment-associated organic

contaminants bioassays. Environmental Toxicity and Chemistry 13: 1315-1329.

Hoke, R.A., G.T. Ankley, A. M. Cotter, T. Goldenstein, P.A. Kosian, G.L. Phipps, and F.M.

VanderMeiden. 1994. Evaluation of EqP theory for predicting acute toxicity of fieldcollected sediments contaminated with DDT, DDE and DDD to the amphipod Hyalella

azteca. Environmental Toxicity and Chemistry 13: 157-166.

Howard, D.E. and R. D. Evans. 1993. Acid-volatile sulfide (AVS) in a seasonally anoxic

mesotrophic lake: seasonal and spatial changes in sediment AVS. Environmental Toxicity

and Chemistry 12: 1051-1057.

Jaffé, Rudolf. 1991. Fate of hydrophobic organic pollutants in the aquatic environment: A

review. Environmental Pollution 69: 237- 257.

Johnston, R.K. 1993. Acid volatile sulfides and simultaneously extracted copper, lead, and

zinc in sediments of Sinclair Inlet, Washington. Technical Report 1552. San Diego,

California: Naval Command Control and Ocean Surveillance Center, RDT and E Division.

91 pp.

Kornicker, W.A. and J.W. Morse. 1991. Interactions of divalent cations with the surface of

pyrite. Geochimica et Cosmochimica Acta 55: 2159-2171.

Lake, J.L., N.I. Rubinstein, H. Lee II, C.A. Lake, J. Heltshe, and S. Pavignano. 1990.

Equilibrium partitioning and bioaccumulation of sediment-associated contaminants by

infaunal organisms. Environmental Toxicity and Chemistry 9: 1095-1106.

C-4
AVS Proceedings

Landrum P.F, S.R. Nihart, B.J. Eadie, and W.S. Gardner. 1984. Reverse phase separation

method for determining pollutant binding to Aldrich humic acid and dissolved organic

carbon of natural waters. Environmental Science and Technology 18: 187-192.

Landrum, P.F., H. Lee II, and M.J. Lydy. 1993. Toxicokinetics in aquatic systems: model

comparisons and use in hazard assessment. Environmental Toxicity and Chemistry 11: 1709-

1725.

Long, E.R., D.D. MacDonald, S.L. Smith, and F.D. Calder. 1995. Incidence of adverse

biological effects within ranges of chemical concentrations in marine and estuarine

sediments. Environmental Management 19: 81-97.

Luoma, S.N. and J.L. Carter. 1993. Understanding the toxicity of contaminants in

sediments: beyond the bioassay-based paradigm. Environmental Toxicity and Chemistry

12: 793-796.

Lyons, W.B. and W.F. Fitzgerald. 1981. Trace metal speciation in nearshore anoxic and

suboxic pore water. in C.S Wong, E. Boyle, K.W. Bruland, J.D. Burton, and E.D. Goldberg

eds. Trace Metals in Sea Water, New York: Plenum Press. pp 621-641.

MacDonald, D.D. 1994. Approach to the assessment of sediment quality in Florida

coastal waters. Volume I. Development and evaluation of seven quality assessment

guidelines. Tallahassee, FL: Florida Department of Environmental Protection. 122 pp.

McGinley, P.M., Katz, L.E., and W.J. Weber, Jr. 1993. A Distributed Reactivity Model of r

Sorption by Soils and Sediments. 2. Multicomponent Systems and Competitive Effects.

Environmental Science and Technology 27: 1524-1531.

McGroddy, S.E. and J.W. Farrington. 1994. In situ and Desorption Partition Coefficients

Measured for PAHs and PCBs in Boston Harbor Sediments. Preprinted Extended Abstract,

American Chemical Society 34: 123-126.

Meador, J.P., E. Casillas, C.A. Sloan, and U. Varanasi. 1995. Comparative

bioaccumulation of polycyclic aromatic hydrocarbons from sediment by two infaunal

invertebrates. Marine Ecology Progress Series 123: 107-124.

Morse, J.W. and J.C. Cornwell. 1987. Analysis and distribution of iron sulfide minerals in

recent anoxic marine sediments. Marine Chemistry 22: 55-69.

Morse, J.W., F.J. Millero, J.C. Cornwell, and D. Rickard. 1987. The chemistry of the

hydrogen sulfide and iron sulfide systems in natural waters. Earth-Science Reviews 24: 1-42.

Murphy, E.M., J.M. Zachara, S.C. Smith, J.L. Phillips, and T.W. Wietsma. 1994. Interaction

of Hydrophobic Organic Compounds with Mineral-Bound Humic Substances. Environmental

Science and Technology 28: 1291-1299.

Ortego, Lisa S. and William H. Benson. 1992. Effects of dissolved humic material on the

toxicity of selected pyrethroid insecticides. Environmental Toxicology and Chemistry 11: 261-

265.

Parkerton, T.F., J.P. Connolly, R.V. Thomann, and C.G. Uchrin. 1993. Do aquatic effects or

human health endpoints govern the development of sediment-quality criteria for nonionic

organic chemicals. Environmental Toxicity and Chemistry 12: 507-524.

C-5
AVS Proceedings

Pavlou, S.P. and R.N. Dexter. 1979. Distribution of polychlorinated biphenyls (PCB) in

estuarine ecosystems. Testing the concept of EqP in the marine environment. Environmental

Science and Technology 13: 65-70.

Rebhun, M., R. Kalabo, L. Grossman, J. Manka, and Ch. Rav-Acha. 1992. Sorption of

organics on clay and synthetic humic-clay complexes simulating aquifer processes. Water

Research 26: 79-84.

Schlautman, M.A. and J.J. Morgan. 1993. Effects of aqueous chemistry on the binding of

polycyclic aromatic hydrocarbons by dissolved humic materials. Environmental Science and

Technology 27: 961-969.

Suedel, B.C. and J.H. Rodgers, Jr. 1994. Development of formulated reference sediments for

freshwater and estuarine sediment testing. Environmental Toxicity and Chemistry 13: 1163-

1175.

Swartz, R.C., D.W. Schults, T.H. Dewitt, G.R. Ditsworth, and J.O. Lamberson. 1990.

Toxicity of fluoranthene in sediment to marine amphipods: A test of the EqP approach to

sediment quality criteria. Environmental Toxicity and Chemistry 9: 1071-1080.

Swartz, R.C., P.F. Kemp, D.W. Schults, and J.O. Lamberson. 1988. Effects of mixtures of

sediment contaminants on the marine infaunal amphipod, Rhepoxynius abronius.

Environmental Toxicology and Chemistry 7: 1013-1020.

Tessier, A., Y. Couillard, P.G.C. Campbell, and J.C. Auclair. 1993. Modeling Cd

partitioning in oxic lake sediments and Cd concentrations in the freshwater bivalve

Anodonta grandis. Limnology and Oceanography 38: 1-17.

Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic

chemical accumulation in aquatic food webs with sediment interaction. Environmental

Toxicology and Chemistry 11: 615-629.

U.S. EPA. 1994. Briefing report to the EPA Science Advisory Board on an EqP approach

to predicting metal bioavailability in sediments and the derivation of sediment quality

criteria for metals (EPA-822-D-94-002). Washington D.C.: U.S. EPA, Office of Water

Resource Center.

Van Der Kooij, L.A., D. Van De Meent. C.J. Van Leeuwen, and W.A. Bruggeman. 1991.

Deriving quality criteria for water and sediment from results of aquatic toxicity tests and

product standards: application of the EqP method. Water Research 25: 697-705.

Webster, J. and I. Ridgway. 1994. The application of the EqP approach for establishing

sediment quality criteria at two UK sea disposal and outfall sites. Marine Pollution Bulletin

28: 653-661.

Winger, P.V., P.J. Laiser, and H. Geitner. 1993. Toxicity of sediments and pore water from

Brunswick Estuary, Georgia. Archives: Environmental Contamination and Toxicology

25: 371-376.

Zhuang, Y., H.E. Allen and G. Fu. 1994. Effect of aeration of sediment on Cd binding.

Environmental Toxicology and Chemistry 13: 712-724.

C-6
AVS Proceedings

Appendix D

Acronyms and Abbreviations

ACS American Chemical Society

AET apparent effects threshold

AVS acid volatile sufides

AWQC ambient water quality criteria

BASF bioaccumulation sediment factor

CAS Columbia Aquatic Sciences

Cd cadmium

CERCLA Comprehensive Environmental Response, Compensation, and Liability

Act

CRCB Coastal Resource Coordination Branch

Cd dissolved or pore water phase chemical concentration

Cs solid phase chemical concentration

Cu copper

DOC dissolved organic carbon

EPA Environmental Protection Agency

EqP equilibrium partitioning

ERL effect range low

ERM effect range medium

Fe iron

foc organic carbon concentration

HAZMAT Hazardous Materials Response and Assessment Division

Hg mercury

IW interstititial water

IWTU interstititial water toxic units

KDOC dissolved organic carbon partitioning coefficient

Koc organic carbon partitioning coefficient

Kow octonol-water partitioning coefficient

LC lethal concentration

Me metal

Mn manganese

MSL Marine Sciences Laboratory

D-1
AVS Proceedings

Ni nickel

NMFS National Marine Fisheries Service

NOAA National Oceanic and Atmospheric Administration

OC organic carbon

PAH polycyclic aromatic hydrocarbons

Pb lead

PCB polychlorinated biphenyl

PEL probable effects level

PPTU predicted pore water toxic units

PRP potential responsible party

PTU porewater toxic units

QA/QC qualtity assurance/quality control

RGS reporter gene system

S sulfur

SEM simultaneously extracted metals

SLC screening level concentrations

SOP standing operating procedure

SQC sediment quality criteria

TBT tributyltin

TIE toxicity identification evaluation

TEL threshold effects level

TOC total organic carbon

TU toxic unit

USACOE U.S. Army Corps of Engineers

USFWS U.S. Fish and Wildlife Service

WES Waterways Experimental Station

Zn zinc

D-2