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Office Of Water 

(WH-556) 

EPA 

United States 

Environmental Protection 

Agency 

EPA 823-R-92-006 

September 1992 

Sediment Classification 

Methods Compendium 
SEDIMENT CLASSIFICATION METHODS 

COMPENDIUM 

Prepared by 

U.S. Environmental Protection Agency 

Sediment Oversight Technical Committee 

EPA Work Assignment Managers 

Beverly Baker and Michael Kravitz 

Office of Science and Technology 

Washington, DC 20460 
ACKNOWLEDGMENTS 

This document was prepared by the U.S. Environmental Protection Agency Sediment Oversight 

Technical Committee. The Sediment Oversight Technical Committee, chaired by Dr. Elizabeth 

Southerland of the Office of Science and Technology, has representation from a number of 

Program Offices in Headquarters and the Regions. 

Appreciation is extended to the authors of each chapter contained in this document. Critical 

reviews of portions of the document were provided by the following persons: G. Allen Burton, 

Jr., Tom Chase, Rick Fox, Audrey Massa, George Schupp, and Howard Zar. 

Assistance in preparation and production of the Compendium was provided under EPA Contract 

No. 68-C8-0062. 
CONTENTS 

Chapter One: 

Chapter Two: 

Chapter Three: 

Chapter Four: 

Chapter Five: 

Chapter Six: 

Chapter Seven: 

Chapter Eight: 

Chapter Nine: 

Chapter Ten: 

Chapter Eleven: 

Chapter Twelve: 

Chapter Thirteen: 

Chapter Fourteen: 

Introduction 

Quality Assurance/Quality Control, Sampling, and Analytical 

Considerations 

Bulk Sediment Toxicity Test Approach 

Spiked-Sediment Toxicity Test Approach 

Intersititial Water Toxicity Identification Evaluation Approach 

Equilibrium Partitioning Approach 

Tissue Residue Approach 

Freshwater Benthic Macroinvertebrate Community Structure and 

Function 

Marine Benthic Community Structure Assessment 

Sediment Quality Triad Approach 

Apparent Effects Threshold Approach 

A Summary of the Sediment Assessment Strategy Recommended by 

the International Joint Commission 

Summary of Sediment-Testing Approach Used for Ocean Disposal 

National Status and Trends Program Approach 
CHAPTER 1 

Introduction 

1.1 BACKGROUND 

The problem of contaminated sediments is 

widespread in freshwater and marine systems 

throughout the world. Contaminated bottom 

sediments can have direct adverse impacts on 

bottom fauna. Contaminated sediments can also 

be a long-term source of toxic substances to the 

environment and can impact wildlife and humans 

through the consumption of food or water or 

through direct contact. These impacts may be 

present even though the overlying water meets 

water quality criteria. As a result, something 

more than the traditional water and effluent 

quality-based control and monitoring approaches 

will be needed to protect and restore the quality 

of the Nation’s rivers, lakes, estuaries, and 

embayments. 

In recognition of the significance of the 

problem, the U.S. Environmental Protection 

Agency (EPA) has begun a comprehensive contaminated sediment program. The effort began in 

1985, when EPA examined the potential national 

extent of sediment contamination using existing 

sediment monitoring data from the EPA Storage 

and Retrieval System (STORET) database (Bolton 

et al., 1985). These data were compared to 

organic carbon-normalized threshold concentrations calculated from existing water quality 

criteria using the equilibrium partitioning model. 

In 1986, the EPA formed the Sediment Criteria 

Technical Advisory Committee to examine possible approaches for deriving regulatory criteria for 

sediments. In 1988, EPA formed two oversight 

committees to take a comprehensive look at the 

whole range of contaminated sediment issues: the 

Sediment Oversight Steering Committee, which is 

responsible for overall management of the program, and the Sediment Oversight Technical 

Committee, which is oriented toward technical 

issues and is the implementation arm of the 

Steering Committee. These committees have 

prepared a draft outline describing EPA’s Contaminated Sediment Management Strategy and have 

formed working groups to focus on specific issues 

and approaches to sediment management. The 

committees are also sponsoring a number of 

activities aimed at providing basic information 

about contaminated sediment issues to persons 

within the Agency and to the interested public. 

This compendium of sediment assessment methods 

is one of the committees’ products. 

An important initial step in addressing the 

contaminated sediments problem is the identification of scientifically sound methods that can 

be used to assess whether and to what extent sediments are “contaminated” or have the potential for 

posing a threat to the environment. The Sediment 

Oversight Technical Committee compiled this 

compendium of sediment assessment methods 

through the efforts of the committee members and 

others who are experienced in the state of the art 

in sediment assessment. 

Many factors can affect the kinds and magnitudes of impacts that contaminated sediments have 

on the environment. The sediment assessment 

tools vary in their suitability and sensitivity for 

detecting these different endpoints and effects. It 

is, therefore, important to properly match the 

assessment methods to the site- and programspecific objectives of the study being conducted. 

The suite of assessment methods presented in this 

compendium offers a rich repertoire of tools from 

which to select the most suitable tests for a given 

situation. 

Unfortunately, there simply is no single 

method that will measure all contaminated sediment impacts at all times and to all biological 

organisms. This is the result of a number of 

factors, including environmental heterogeneity and 

associated sampling problems, variability in the 

laboratory exposures, analytical variability, differing sensitivities of different organisms to different 

types of contaminants, the confounding effects 

caused by the presence of unmeasured contaminants, the synergistic and antagonistic effects of 

contaminants, and the physical properties of 

sediments. While one method will suffice for 
Sediment Classification Methods Compendium 

some circumstances, itis often advisable to use 

several complementary methods rather than a 

single one. When several of these approaches are 

used together, they can provide additional insights 

into the nature and degree of sediment contamination problems. The use of complementary assessment methods can provide a kind of independent 

verification of the degree of sediment contamination if the conclusions of the different approaches 

agree. If the conclusions differ, that difference 

indicates a need for caution in interpreting the 

data since some unusual site-specific circumstances may be at work. The importance of this type 

of verification increases with the significance of 

the decisions that must be made using the information obtained. In fact, the actual decisionmaking frameworks within which the compendium 

methods are used often include this verification in 

the concept of tiered testing. 

The assessment methods presented in the 

compendium are continually being refined and 

improved. Additional methods are also being 

developed. As these methods are developed and 

verified, they will be incorporated into future 

updates of the compendium. 

1.2 OBJECTIVE 

This document is a compendium of scientifically valid and accepted methods that can be used 

to assess sediment quality and predict ecological 

impacts. 

Some regulations require the use of certain 

types of tests (e.g., the Toxicity Characteristic 

Leaching Procedure under the Resource Conservation and Recovery Act), criteria (e.g., the limitations in the London Dumping Convention), and 

procedures (e.g., risk assessment under the Comprehensive Environmental Response, Compensation, and Liability Act). Additional guidance may 

be issued in the future to provide direction when 

addressing sediment contamination under particular regulatory programs including these, or other, 

required tests and approaches. These other test 

procedures will not be presented in this compendium, however, because the intent here is to provide 

the most useful overall measures or predictors of 

ecological impacts currently in use rather than 

procedures that may have limited application 

outside of a particular regulatory framework 

Nevertheless, many of the methods presented in 

the compendium can be used as part of regulatory 

and/or remedial actions. 

Guidance on how to use the compendium 

methods in a decision-making framework will be 

provided in forthcoming documents and will likely 

include both chemical and biological methods in 

a tiered hierarchical framework suitable for testing 

various hypotheses and endpoints. Currently such 

a document has been prepared by the Sediment 

Oversight Technical Committee to summarize 

existing EPA decision-making processes for 

managing contaminated sediments (Managing 

Contaminated Sediments: EPA Decision-Making 

Processes, USEPA, 1590). The information 

provided in the compendium on the relative 

strengths and weaknesses of the different assessment methods can provide assistance in selecting 

the appropriate methods. 

1.3 OVERVIEW 

The compendium is organized in the following 

manner. The remainder of this chapter gives a 

broad overview of the assessment methods in the 

compendium. The information is presented in 

tabular form to facilitate comparisons between the 

different methods. Chapter 2 outlines quality 

assurance/quality control, sampling, and analytical 

considerations that apply to all of the methods. 

Method-specific information is also provided 

where the procedures differ from the general ones. 

The remaining chapters give specific information on each of the sediment assessment methods. 

The information is organized in a consistent 

manner for each assessment method so the reader 

can readily compare the relative strengths, weaknesses, and applicability of each method in order 

to select the best method(s) for a specific situation. The information provided for each method 

includes the following: 

1-2 
l How each method is currently used or 

could be used; 

l A detailed description of the method, 

including types of data, equipment, and 

sampling procedures needed; 

l The applicability of the method to the 

protection of wildlife and humans; 

l The utility of the method to produce 

numeric sediment quality criteria; 

l The method’s applicability to making 

different types of sediment management 

decisions; 

l The method’s advantages, limitations, 

costs, level of acceptance, and accuracy; 

l The degree to which the method is actually being used now; 

m How well it is validated; and 

l Its potential future uses. 

Extensive references are provided after each 

method in case any additional details are required. 

The names, addresses, and telephone numbers of 

the authors of the descriptions of each method are 

provided to facilitate additional follow-up. Given 

the limited level of detail in the compendium, use 

of these references is suggested for actual implementation of the methodologies. 

The 12 sediment assessment methods described in the compendium are summarized in 

Table 1-1. The assessment methods can be 

categorized in many different ways. Differentiation could be made between numeric methods and 

descriptive methods. Numeric methods are chemical-specific and can be used to generate numerical 

sediment quality criteria (SQC) on a chemical-bychemical basis. A potential drawback of descriptive methods is that they are not chemical-specific 

and cannot be used alone to generate numerical 

sediment quality criteria for particular chemicals. 

On the other hand, descriptive methods can be 

used to directly assess the overall impact of all 

chemicals that may be present in a sediment, 

whereas it is difficult to use the chemical-specific 

methods to predict the combined effects of several 

chemicals. 

Another differentiation that is often made 

among different sediment assessment methods is 

whether they are based on the measurement of the 

concentrations of chemicals of concern or on the 

measurement of biological impacts. For methods 

that have ecological validity, this differentiation 

really applies only to the practical implementation 

of the methods rather than to their scientific basis 

since al1 ecologically valid methods must ultimately be based on an ability to predict or measure 

biological effects. Many of the assessment methods use both chemical and biological testing or 

observation. 

Yet another differentiating factor is whether 

the method uses interstitial water (pore water), 

elutriate, or bulk sediment (whole, including the 

solids and interstitial water). This difference also 

relates primarily to implementation rather than to 

a substantive scientific difference since the chemistry of interstitial water and that of the bulk 

sediment are closely linked Except for cuntaminants that might be transferred directly by ingestion, interstitial water is the medium through 

which the contaminants in the bulk sediment are 

transferred to the affected organisms. 

Some of the assessment methods (which 

would be more accurately characterized as approaches) described in the compendium combine 

numeric and descriptive measures. For example, 

the Sediment Quality Triad (Triad) and Apparent 

Effects Threshold (AET) approaches employ bulk 

sediment toxicity testing, benthic community 

structure analysis, and concentrations of sediment 

contaminants. The Triad is both descriptive and 

numeric, depending on its use. Typically, the 

Triad approach has been used in a descriptive 

manner to identify contaminated sediments. It has 

also been used, however, to generate criteria for 

several chemical contaminants. The International 

Joint Commission (DC) approach would be more 

accurately described as an assessment strategy 

since it employs several of the other sediment 

assessment methods in a tiered, comprehensive 

13 
Sediment Classjfication Methods Compendium 

Table 1-l. Some Characteristics of the Sediment Assessment Methods. 

or quality criteria a 

correlations, and human 

mc cemmun 

M commuruty structure are 

ment chemistry, toxicity, nd 

lment concentrahon of a contaminsnt &ove which sWstkally signm 

inf:a) would ahvays be expeded. AET values ue empbkally derived from paired fidd d&a tar 

hit community structure and cu~cerlrdins of cotiaminants in sediments and 

bidogicaI tissues and (2) a ddaibd assessment thd Is based on B phased sampling of the 

bored testing strsdsgy consisting of physical, ohsmical, urd biilogii testing to pr 

and water column impacts of dredged eedirnsnt disc. 

Trends Program 

14 
procedure. The Sediment-Testing Approach Used 

for Ocean Disposal is the tiered, comprehensive testing procedure developed by EPA and the 

U.S. Army Corps of Engineers (USAGE) for 

determining the suitability of dredged material for 

disposal at designated disposal sites. The procedure is specified in Evaluation ofDredged Material Proposed for Ocean DQosaZ-Testing Manual, 

commonly referred to as the 1991 Green Book 

(USEPA/USACE, 1991). 

To facilitate the user’s selection of the most 

suitable sediment assessment method, Tables l-2 

through l-5 highlight the major characteristics of 

each method. Information from individua1 chapters that is useful in management decisions is 

presented in summary form and includes method 

descriptions and uses, data and sampling required, 

ability to generate numerical sediment quality 

criteria, and outlook for future use. More pointedly, the reader will learn what each method predicts, what it assumes, how much it will cost, and 

why one might choose a particular method over 

another for a specific situation. 

Regardless of which of the compendium 

methods one uses, several considerations must be 

addressed: a sampling program needs to be 

designed; samples need to be collected, stared, 

and analyzed, and quality assuiance/quality control 

is needed throughout the process to determine the 

uncertainty associated with the results of the 

assessment. Sampling design and QAIQC issues 

will be discussed in Chapter 2. 

1.4 REFERENCES 

Bolton, S.H., RJ. Breteler, B.W. Vigon, JA. 

Scanlon, and S.L Clark. 1985. National 

perspective on sediment quality. 

USEPA. 1990. Managing contaminated aedimen& EPA decision-making processes. U.S. 

Environmental Protection Agency, Sediment 

Oversight Technical Committee. EPA 506/6- 

9oPo2. 

USEPA/USACE. 1991. Evaluation of dredged 

material proposed for ocean disposal-Testing 

manual. U.S. Environmental Protection Agency and U.S. Army Corps of Engineers. 

l-5 
Sediment Cfass+catbn Methods Compendium 

Table l-2. Summary of Sediment Methods and Applications. 

Spiked Sediment 

Toxicity Test 

n resaar& state. 

chemicJ mixtures. 

InterstltW Water 

Toxklty Test and freshwater. 

Fbld-ahcbd bulk 

with TIE procMure5. data, pwtkululy h combi- aadhnenb. 

natlon with other sediment 

&Mkatbnmahods. 

Equillbrlum PartMbnlng Regulatory uses of Equilibrium Yes. Interim SOC for EqP-based SQC hm a Humn hadth, guatic sadhlmnt ohmmiutry, total 

Partitbning (EqP)-based SQC scma chemkals hava posslbb major role in tha Ma, and wkUb. organic arbul 

under devabpment. been dewloped. idenmrcatiotl, monttoring, con~atkn. 

and cbanup of ccntarninated sIba. 

Tbnua Rasklua 

Freshwatw Benthk 

Communtty Strwtura 

and Function 

+na use in remedial and regu- Yes. Most l pplkabb Wll providae~ Humm he&h, 4untk sadbmrltchmn~md 

latwy actiwrr. for nonbnic organic maa8um of ‘affacuw life, and wlldllfo. physka! chuactarlstkB. 

and organomatatik sxposura dcse.’ Wota umpling for rwldw 

omnpcunds. analysb. 

A numbw of uses, including the No. will be mo6t wcco8sful in DlrecUy l pplkabk to sadimantaJfbcuonuning 

astabtiahmmtof~rlaand amcartwithwdknsnt aquatkllfmandunna agrdrunpkr. 

standuds. *bfvd- wMf0, and indirsetly to 

~SUlb. human-ndti 

wiMnfo. 

Marh&nthbcommunHy Dwcrlbes reference conditions, Not alone. Integral Potential for kbnUfying arsctly*bJ@tfJ su!imaltcdbctkmuslng 

Struchr~ baselins conditions. and effecb camponent of Al3 specks that are lndicaUw aquatic Rk mnd uorw a grab or wro sampler. 

of natural and mthropogenk and Sadlmmt of wdimmt contaminmta wildHfa, md klhctty b 

dbtuti. Qualny Triad. at varbus concmtratbw. human h#llh nd &a 

Wlldlk. 
Z-Intrvduction 

Table 1-2. Summary of Sediment Methods and Applications. (Continued) 

Sedlmant Method Current Uw 

Sediment Quality Triad Det5rmin55 extent of pollutioninduced degradation. 

D5termines numerical SOC. 

Spparent Eflsch Threshold Used by several program5 to 

devabp guidelines for protectIon 

of aquatic Iifs in Puget Sound. 

ntarnationai Joint 

~mmbsion 

Intended as guidance for 

a55555ment of contaminabd 

sedimenb in the Great Lakes. 

1991 Green Book Gukianca for dmdglng 5ppllcant5, 

scbnUsts, and regulators. 

Uatbnai Status and Trends 

%ogram Approach 

Initially used to dwdop informal 

guk,biinm for use by the NSLIT 

Program. 

fes. Used for bad, Identifying problem areas, 

‘Al-is, and PCBS. prioritizing and ranking 

degraded areas, and 

predicting whsre degradation will occur. 

Identifying probbm amss, 

identifying problem chemicals in sediments, and 

focusing cleanup actMties. 

Screening sediment5 in 

regulatory programs. 

(55. Evaluation of Areas of 

Concern. Possibb us5 

outsids of Great bkes 

basin. 

zbM-validadss SOC Will be applied to dredged 

wrrsntfy under de- material bvrluations for the 

relopment foreseoabie future. 

kbntify toxic chamkds in 

sediments; rank and prioritize arws for further study: 

a55055 pot5ntial ecological 

hazard5 of contaminated 

sediments; design spiked 

sediment bba5s5ys; 

describs t&c sftecb 

a55odated with a3rtain 

chemical aonc5MaUons; 

quantify likelihood of 

toxicity for range of 

chemical mnc5ntraUon5. 

Protut Human Hrfth, 

Aqu5tlc Lib, wtd/w 

WIldlIfe 

Typo of Samphg 

Roquir5d 

Aquatic life directly, 

wildlife and human 

health directly and 

Indirectly. 

Fiild-culbct5d 55diment 

Fhm Rdd replicats bemthk 

samples mcornmended. 

Aquatic Ilf5. Fbklcolbcted sediments 

from50station5ormor5 

recemmendsd. Conduct 

ch5mkal t5st5 for a wide 

range of ch5mical 

CklSSOS. 

Dirwtiy to aquatk life, Bulk 55dimant colbction, 

and indimctly to wiklllfe benthk community 

and human health. structura, fish au~tamfnant 

body inlrrJM5, l d axtarnal &nommiMas. 

oirectiy to aquatic m, 

wildlife, and human 

hwlth. 

Compr5hen5lv5 samplh7g 

plan for 5edlmlMlt and 

watbr. 

Th5se guMdine piovide Mlnimiz55 th5 n55d for 

an estim5t5 of effects on l dditbnal sampling 

benthkllfe. Theyam through th5 us5 of etistnot intencbd to b5 used ing data. 

for the protectlorl of 

human IHS or wikllits. 

l-7 
Sedimcrrt CILzssjcjc(ltion Metes Compendium 

Table 13. Summary of Sediment Methods and Suitability. 

All classes and 

men& 

;piksd Sedl- Adding (‘spiking9 Test, retsrenc5, hY type. All da5555 and In tieory, can us5 Can be used in Canbeusedin Can dstsnnitm 

nent Toxicity 55diment5 wilh and conlroi sadi- combinations of any organism. dev&ping combination with toxldty pdor b 

r-t one chemial or a ment data. Physi- ch5mk.ai5. critaria. can wastelslolld alloca- dbposal. 

mbdurs. cai, chemical, and identify 5xbnt of tkm models to 

bblogknl data. problem. monitor sstablbh maxitrends, and 55t mum dbwabie 

target cl5anup efitwnt concangoab. tr5tiofl5. 

ntar5Utiai Wa- Porswat5rprepa- Physical, Any typa from Watw-solubb Predkts imp&5 IdOnUff55 ld5al for point Can d-in5 

or Toxiclty ration, tooxlcity chemkal and which 5d5quats nonbnic organ& on organbms sediment soutw controb toxbity prbr b 

kit tests, md TIE bblogkal quantities of por5 catbnk metals, onca toxkant toxksntamdcan and tzimtmihbie dbposd. 

proc5dur55. re5pome, kbntifi- water an I241 and ammonia, respomibb for design nonpoirlt 5ourw5. 

catbn of bxlc ObtdtWd. and thdr lntsrac- to&&y is remediafbn pi5ns. 

ampound5. tknls. idMtMed. Routine 

monltoring. 

Pr5dMs dlemld sulk 5edimMt After Madificatbm of Can pr5dkt toxic Can identffy Radkts concm- Suttabb for adconcentratbn in anatysb and con- d5v&pment of method exbt for dfwta for a range source45 of tratlon5 of a cite- dressing aquatk 

lnter5~ial wat5r centraUon cd total test, tmpctad 8~ diffsrsnt dassm of r5pm5eJnwhm contaminatlon~ and mill abow dbposai. UnsuRand ccmpar55 it b orgmnic carbon. apply to wide of ch5micai5. organisms. id5ntify targ5t whkh ad-ma rbb for addrssschronk wabr rang5 of cleanup l5v5ls. impacts am ing upland 

quality aMsrla. 55dimsnt typ5~5. Iikdy. disposal sibs. 

z-8 
Table 13. Summary of Sediment Methods and Suitability. (Continued) 

Provkiw estab- Sultabb for use 

to residues, and 

link5 chsmical 

bbrcwnulatbn 

isms to sediment aquatic organbms 

chemkby. for reddua. wm~min5Uon. 

Fleshwrbsr Fbki aunmy. col- Varbs from a list Any type. but Many indivfdud Facillt~~s use of Can be used to Extsmhmly used Advised br arem 

Bbnthic l5cUon. 5cdrtg. of famllie5 of tax5 only 5imilar typ55 ct3smkals and benthk macrdn- screen for poten- fw source suhmble )or openCommunity and id5Mkatbn pwent to sp- shoukl ba mm- cla5545 of dwnl- vwtebratss as tiai 5ource5 of charactsrization lake dbpceal. 

Structure of benulk organ- cieSk~l - PMd. cab. indkata organ- cOrlhh5tiUl. and conti. 

lnd Function i5m5. my md 5num5ra- isms. 

tbns. 

Muhe Bsnthic Cdidon, sorting, Numb of taxa, Anytyps, but Applbs to gwwral Fadlttat55 use of Has not bwn Llmtbdwluein Not required in 

tbmmunity and kiadicatkn abund- of only similar types mtegarios wtth bar&k -dn- ulwdbwt spedk sourrm te&lg of wdistnJctura of banthic orgm- each taxon. and stmuld bs mm- 5xepUons bassci wrt5brat55 as in- sediment qudity cfw5cterlzation. ment to k 

lams. bii55 and oon- par5d. on bvd of organ- dicator organbms. criteria for pollubd drbdg5d under 

willbnai sedi- k bnrkhment. Rtwarch b ~~cKI- marine sedhnents. 55dcm5 401 and 

ment chunbtry 5dtopnYdktsp5 404ofth5ckan 

vuidhs. dfkefbclson WIltsrACt. 

pot5nUal predators. 

sodimall U505 sediment Sediment chemis- Any type. -All chemkeh and All bbkgkal sf- A comprshsnsiw Cornpreh5nsiw i-fl5tory of rsguiaoUalUy Trfd ihmistry, 55dL try, mdhlmt bx- dasws. fads data bawd approach Umt and oompbmanb tory uaa. 

msnt bbass5ys, kity, and bsnthk on l singb ape- allovm for all pate- TlE prugwns for 

and fn SHV Mdogi- infauna data. tbs. nlal Int5racUons fdkmlts. 

cd variabbs. b5tw55n chemkd 

mbdures and the 

emironment. 

1-9 
Table l-3. Summary of Sediment Methods and Suitability. (Continued) 

lemaraa5and dktwlwhrdbs determined. 

remedial actkn. 

rekr5nca conditorbua, chenlkal 

amcanlr~m in assa55mantofin4ph- a--?- tura, md external 

ment. l tmormaiitb5. 

1991 Qremn Tbral-lmlng Phydaiandche- Anytypawiul wkierangaof cbmidar5afbcb owdopadb Not irlbnded u5edindd5kn3ook Pr0-dur-b mkai senllment exwptbn of ax- organk md Moor- on marine organ- determina watsrcharahth data. Bloassay, tmmdyaumaa gmkchwnkab. l5m5t5pr~5~- c&mm UKI -dbpoul 

dredged mahwlal bioaaxmuiation, Mguiu-grain 

andpradktR. ard nekl5p5de5 5edimerlb. 

impact Induda data 

physkal md dlomlml ssdhwnl 

evaluatkm, bxklty, and bkrxumulafbn shrdlr. 

talk of orgmbms benthk UmMng 

indigenous to Pamli55ibb 

-MM Concenbatlar 

Matariai Di5po5d (LPC) ComPii~ 

sibs. for dredged 

ma83M. 

l-10 
I-hztn3ductian 

Table 13. Summary of Sediment Methods and Suitability. (Contmued) 

data from laboraand tidd studies. 
Table 14. Summary of Sediment Methods and Ease of Use. 

Easily Inbwpratabls. BlobgIal data subjected b “passfall’ 01 some explanation. 

ntadtlal Wabr 

r0ddty Teat 

strdghtfolward -MC. A!l8OflSltlvSOr~ sound fhaoratlatl Field aampllng, porn war Rewlts edy hrbrpretbd. 

-b- pl- con-m basis. wwdh ~XlcHy t-M 

highly wnslthm n--w. and TIE prowdufe8. 

instrumentatbn. 

cakulatbnm ua oapwldMtcrloo8td Loveb of protacuon wide aax+la. Fbqulm kuprowbn but 

stralghtfcrwd wllh a4bdlng rltwpedllc of sot slmlkr b “iitgLlz*~ prwldes putlnant lnformanawswy da& chemlcd data. thOSOdWatW tbn. 

qudlty rYltarb - 

deemed protwtk 

of 95% of crgnbmr. 

rbaua ReddIm stralgMtomud. Cd of SO0 9wtordly Doeanottendtobe Accaptedasa Varlaafromnonablargo. Varh3sbynumherndnatura 

l~klWuuh/llclr dhar comandn bask for regulatory of amtamlnatbn, complexity 

mb. or Wberd. decbbnr. of dbtributbn, and regulatory 

applbatbn. 

Fra&water Ekmthk Equlpinent and c700 pm Sam* db. High. wdeaocepfanw R+3ub can be ganamtod Oafa lntwpretatbn requires 

ConwnMy SWuctum matarlds Inox- from a hblortcal wlthln 1 day. an 0xpeR RalxIlbJ arm eadly 

mnd Functhm pemlw and P-m. Inwrporabdtntoa 

minlmal. Orpk management strategy. 

bmsdlffk4ttosort 

and #mtlfy. 

f-12 
Table 14. Summary of Sediment Methods and Ease of Use. (Continued) 

Wde acceptancs Fiild effort. laboratcfy 

from a hbtoriil work, and analysb may 

take saveral months lo a 

Apparent Effects 

Threshold 

data canparbons. 

Trends Rogfmm 

Subwquenl use 

database b relatlwfy 

l-13 
Table l-6. Summary of Sedlment Methods and Extent of Use. 

roxrctty Test 

Some fbkl validation; more l Prombing for direct maasumnerrt 

hasb needed on maamental ccndttions. 

n@rstltM Water 

r0* T-t 

All sediment types wld High. Widaacaptamwfu Little Md validatlan; mare b ’ Extrandy promkIng: on4y meulod 

l vlronmMtd condltknr. freshwater and marina nacesswy. that dlractly Indudes #lo ldenufkaapplbaucmr. tbnofcmlpoundmrnpomlbtafor 

wxid(y. 

l Further dewbpment noedad. 

EqP-b&wd SOC apply to Each EqP-l!ased SOC underEPArwbwtw someftddvand~ mom l ontyprocedlNafuduhfatlonof 

sadiments with greater WllhWOIB- ragulamy uses d EqP- II newssuy. SO0 thmt b generk across wllthan 0.2% organic car- degree of urwrtalnty. bawd sot. manb, accomb la bbadWltty. 

bcnandnonkmkdwm- and rdmtaa mfbcb b dmmiub. 

lcahforwhlchcrbda 

are l vdlabb. 

nsaw Remktw uttypes. Generally high. Widaaoxptanca. Some flaki vatklaUon for l Can k Impbmanbd wtfh mlnlml 

Mvktual chemk&, mm eflort 

for chemiul mlxtwer. l cmtrddabbmashotibadavdm. 

l Fb~valkl~ofrealdw-brued 

-w-l-- 

eMMtld. 

l-14 
Table 1-5. Summary of Sediment Methods and Extent of Use. (Continued) 

. 

. 

. 

; l 

. 

. 

. 

. 

. 

Accur8cy l d 

Prrclslon I Ext8mof II Sediment Method Extwrt of Us. Fkld Wld4on 

Envlronmsnbl 

Appllcrbltlty 

Lotk and lentk freshwater ecosystems. 

OuUook for Futun UBO 

Outlook good because benthk 

maaoinvsrfpbrates prod& 

substantial infwmatkn that 

chemkal and tmdcity data abns 

cannot provide. 

DevulopmMl most needed h ambining benthk community msessments with chemkal and toxkologi. 

cal data. 

II Freshwater Benthii 

Community 

High. Wie acwptma. An in SiftI study; therefor8, 

cc~sbtently and uccurately 

~tructurs assesses 0nvironmenW 

and Function 

Direct measure of envi7 

rmnmtal effects. 

High; If neetssary replicates are obtained. 

Outbok bright with continued 

devekpmmt toward now data 

analysis mhuds to reduce cost or 

varlabHlty wlthln data. 

Marine Benthic 

Community Structure 

Valued tool for several Am h sihr study; themtire, 

decades. conshtentiyandaaaratdy 

assesses em&onmental 

ww. 

Sediment Quality Exfmmsly high. 

Triad 

Not qwntitnUvely determind;sq3actedtobe 

high. 

High potMtld. Provld88 obj8cuv8 

lntwmatlon to judge e*t of pollutlon-lrxluad d8gradatbn. 

khdhod devekqrnent and standard. 

kntbn mamary. 

Rocantly dsvebp8d. E3y It, nature an h s&4 study; 

Has been used lo Mul- thersfore, wtomatlcatly fbldHfy degrnded Ivens. valid&d. 

Apparent Effects 

Thrr-hold 

Sensitive and effident 

The number of statlons 

usedhnsamuk8dedf8c 

on AET uncertainty. 

Used by Puget Sound FkM-vdidnti for Puget High potentld for mgbnd use. 

8g8ndss for mgul8tory Sound. Furthw testing 

guldellms. Abo wlddy desired before applkation of 

used by othem. hget Sound AETs b other 

guogrnphk regions. 

Pubtishal in 1999. 

lndlviduml methoda 

wldety used and 

-=pbsd. 

Fimt fldd vdMatlon ln 1999. 

1991 as put of EPA ARCS 

program. 

Guklancn wtll be applied Large portkns were fbldto l N wdu8tkns for vdklated In the past addidredged nubrinl thrt b tional pojscts planned. 

~loposed for dbw=l 

outsld8ofth8b8sdin8of 

th8 twrltorial WM. 

Intnmathal Joint 

Commhsic i 

High. Not quantitathm~ determined. Eqwxted to be 

high. 

Potential for wldsspraad use In 

Great m basin and dsewha*. 

Strongly supports oxbmh-8 CM progrcvn. 

EPA and USACE contlrnm to supporttlmgt.khcanrtknallyand 

WOfW. 

Ongohg public and ptvab research and dsvdopment wifh amcomitant documMt updncs. 
Table 14. Sumwwy of Sediment Methods and Extent of Use. (Continued) 

Envlronmwttal 

Natbnd staaJs and Highly applicable to the Once the minimum l-k4SbWflUSedby VaHdaUons hava not yet l oulbokbgoocl. slrlmula 

Trends Rogram number of data sets b approach relies on existing da& 

environmental data. determined to devdop and Trends Program, ouler regknlapedfk gukblines 

could be emslly devolm udng 

variebility Is minimal. and the Fbrkla Depart- region-spedfkz data. 

Alxuracy In predicting ment of EnvIronmental l Approndranbeusedbntklab 

toxkii has not been Rsgulatlcm. A variation critala determined wlth other 

of the approwh Isbeing slngle-methcd appwchas. 

developed by the C& - !5evemltypesofdlfaaferlealed 

fornia Waler to futlhar dweklp me npprondl. 

l-16 
CHAPTER 2 

Quality Assurance/Quality Control, 

Sampling, and Analytical Considerations 

The purpose of this chapter is to provide a 

brief introduction to some of the most important 

terms and concepts that are integral to the design 

of an adequate program for sediment sample 

collection, handling, and analysis. This chapter is 

intended only as a general guide to sediment 

sampling and should not be used as an instruction 

manual for collecting samples. The subjects 

mentioned will not be dealt with in an exhaustive 

manner. The reader is referred to the references 

cited in this chapter for more complete guidance 

on the particular techniques. 

2.1 ESTABLISHING DATA QUALITY 

OBJECTIVES 

Fundamental to the process of designing a 

study is the establishment of data quality objectives (DQOs). The most carefully collected and 

analyzed data are of no use if the data collected 

are insufficient or of the wrong type. To avoid 

either of these and other potentially costly errors, 

EPA has initiated the use of the DQO Process. 

The DQO Process is a management tool designed 

to help data users and data collectors design the 

best sampling strategy to reach their objectives 

while minimizing resource requirements. It is a 

multistep, systematic approach to data collection 

that enables the manager to refine goals and 

objectives and help answer the question, “How 

much data is enough?” As the steps of the DQO 

Process are followed, the decisions made in 

previous steps should be reviewed to ensure 

consistency and cohesiveness. 

The first step in the process is to specify the 

problem and identify limitations of time or resources on the data-collection effort. This process 

allows one to evaluate his or her current knowledge base of the problems and identify all available resources. The next step is to identify what 

decisions or activities will be made based on the 

data. The answer to this question is vital to 

ensure the collection of the right type of data. 

The decision goals should be as narrow in scope 

as possible, and considerable effort may be required to define them properly. 

The third step involves identifying all variables needed to make a decision. This step 

focuses on eliminating the potential measurement 

or collection of data that may not actually be used 

in the decision-making process. The next step 

requires the data collector to set or define the 

boundaries of the study, including the population, 

which could consist of people, objects, or media, 

and the boundaries on the population, including 

space, time, and area. 

Developing a decision rule, or how the data 

will be used and summarized, is the next step in 

the process. This step involves describing how 

the study results will be compiled or calculated 

and defining the decision rule in an “If . . . , then . ..” 

format. The statement should incorporate the 

study results as “If the results are this, then the 

action should be this.” For example, “If PCB 

levels in fish are greater than 2 ppm, then a fish 

consumption advisory will be issued.” This step, 

along with the others, helps define the data collection effort by identifying the data needed to fulfill 

the decision rule. 

A very important step in the DQO Process is 

specifying the limits of uncertainty acceptable in 

the data. These limits can be expressed as acceptable false-positive and false-negative error rates 

for the decision. These error rates must be based 

on careful consideration of the consequences of 

incorrect conclusions being drawn from the data. 

The definitions of false-positive and false-negative 

errors vary with the decision being defined. If a 

decision to take regulatory action is being made, 

a possible false-negative error could result in no 

action being taken because incorrect data results 

indicated there was no problem. The opposite 

could also occur, where a false positive error 
Sediment Classification Methods Compendium 

results in regulatory action being taken when no 

problem exists. It is essential that the potential 

consequences to economic, health, ecological, 

political, and social issues be considered when 

deciding on acceptable false-positive and falsenegative error rates. This step may involve the 

consultation of a qualified statistician. 

Finally, all steps in the DQO Process should 

be reviewed to design the most efficient sampling 

study. Considerations including cost, time, defined boundaries, the decision rule, and all other 

factors defined and specified during the DQO 

Process should be incorporated. 

One can refer to “Planning Issues for Superfund Site Remediation” in Hazardous Material 

Control (Ryti and Neptune, 1991) for an excellent 

example of applying the DQO Process to an actual 

situation. 

Quality assurance and quality control are 

integral components of every aspect of a program’s activities. The collection of reliable data 

is contingent on the use of and adherence to a 

good Quality Assurance Project Plan; the development of a sound sampling study is contingent 

on the use of the DQO Process; and use and 

implementation of the DQO Process is contingent 

on a Quality Assurance Program Plan. 

2.2 SAMPLING DESIGN 

2.2.1 Test, Reference, and Control Sediments 

In sediment quality evaluations, there is a 

substantial precedent for using comparisons 

between sites rather than comparison of testing 

results to an independently set numerical benchmark. This is the result of a number of factors 

including the standard procedures used in biological testing, the paucity of scientifically acceptable 

numerical sediment quality criteria or standards, 

and the long-standing “nondegradation” philosophy used in evaluating the acceptability of 

dredged material for open-water disposal. The 

degree of sediment contamination in a particular 

area is often evaluated by comparing the structure 

of benthic communities, levels of pollutants, or 

bioassay test results in sediments collected from 

the area being investigated with those. in the 

surrounding area. The terms used to describe the 

different sediments in the comparisons are lest 

sediments, control sediments, and reference 

sediments. 

As used in sediment assays and assessments, 

a test sediment is sampled from the area whose 

quality is being assessed. A control sediment is 

a pristine (or nearly so) sediment, free from 

localized anthropogenic inputs of pollutants with 

contamination present only because of inputs from 

the global spread of pollutants (Lee et al., 1989). 

A control sediment is fully compatible with the 

needs of the organisms used in the assay, is 

known to not cause toxicity, and is used primarily 

to verify the health of the test organisms and the 

acceptability of the test conditions (USEPA/USACE, 1991). The control sediment may be artificially prepared in order to achieve sufficient 

volumes of a known and consistent quality for use 

in standard testing and for culturing test organisms 

(ASTM, 1990). 

A reference sediment, on the other hand, is 

collected from a location that may contain low-to 

moderate levels of pollutants resulting from both 

the global inputs and some localized anthropogenic sources, representing the background levels of 

pollutants in an area (Lee et of., 1989). The 

reference sediment is to be as similar as possible 

to the test sediments in grain size, total organic 

carbon (TOC), and other physical characteristics 

(Lee et al., 1989; USEPA/USACE, 1991; ASTM, 

1990). The physical environment of the reference 

site should also be as similar as possible to that at 

the sites where the test sediments will be collected. This is especially significant for benthic 

community structure comparisons, since community structure can be very significantly affected by 

water depth, physical transport processesuch as 

waves and currents, sediment grain size, and the 

presence of organic debris. 

As used in dredged material assessment, the 

results of assays or evaluations on the test sediments 

are compared to those obtained from reference 

sediments to determine whether the test sediments 

are contaminated. In contrast, the results of assays 

or evaluations using the control sediments are 

usually compared only to past results using those 

2-2 
2-QA/QC, SantpIing, and Analytical Casidtmtions 

same control sediments to ensure that the testing 

was free of some extraneous factors that may have 

affected the reliability of the test. Depending on the 

study objeaives, however, controls can also be used 

as a benchmark against which to compare test 

sediments to de(ermine the relative degree of contamination of sediments collt~ted from different 

sites (ASTM, 1990). 

A clear understanding of the end uses of the 

data is essential in the establishment of an appropriate sampling program. A cost-effective study 

for a qualitative overview of potential contaminated sediment impacts will differ markedly from one 

whose purpose is to make statistically-based 

numerical comparisons with criteria or indexes, or 

to reference sites. 

Sediment sampling programs are most often 

undertaken to achieve one or more of the following objectives: 

B To fulfill a regulatory testing requirement: 

w To determine characteristic ambient levels; 

n To monitor trends in contamination levels; 

w To identify hot spots of contamination; or 

n To screen for potential problems. 

These different objectives will lead to different sampling designs. For example, a study for a 

dredging project may have a specific set of guidelines on sampling frequency, sample site selection 

methodology, and other parameters already determined by existing specific guidance. The design 

for a study to determine ambient levels will strive 

to obtain uniform, random coverage of an area 

through the collection of samples from a relatively 

large cumber of sites. The design for a study to 

track sediment contamination trends will expend 

its resources to sample fewer sites but more often. 

A study to identify hot spots would concentrate 

efforts on fewer sites within zones most likely to 

be contaminated, while an initial screening study 

might take very few, randomly distributed samples 

for analysis together with some “observation” 

samples to supplement the analytical results. 

Available information about the area to be 

sampled and its surroundings should be used in 

determining the final sample design. Knowledge 

about bottom topography, currents, areas of dredging and the frequency of dredging, locations of 

point and nonpoint sources of contaminants, 

distribution of grain sizes, and other factors can 

provide the basis for determining which of the 

sampling designs .to use (e.g., Are there reasons to 

expect localized hot spots of contamination?) and 

where to place sampling locations (e.g., Which 

parts of the area are likely to be similar enough to 

group into the same strata?). Preliminary surveys 

of an area using depth-sounding and sedimentprofiling equipment can prove invaluable in 

delineating vertical and horizontal distributions of 

sediments (IJF; 1988). This information can be 

helpful in planning sediment sampling methods 

(grab samples or core samples) and sample site 

selection (grouping similar areas into strata, 

identifying likely locations of hot spots). 

The methods most often used for selecting the 

sample collection sites are haphazard, worst-case, 

random, stratified random, and exhaustive 

(Higgins, 1988). 

2.2.1.1 Haphazard 

The haphazard method, whereby one selects 

sampling sites based on whim or ease of implementation rather than science or knowledge, really 

reflects the lack of a design. This method has no 

validity and should not be used. 

2.2.1.2 Worst-Case 

The worst-case sampling design is based on 

knowledge regarding the presence and distribution 

of potential sources of sediment contamination in 

an area. II is usually considered cost-effective as 

long as the study objectives are being met. An 

inherent problem with this design is that it results 

in an incomplete characterization of an area and is 

not statistically robust. However, it can be useful 

as an initial survey to determine the potential for 

a contamination problem, which would be fol2-3 
Sediment ClassjCication Methods Compendium 

lowed up with more complete sampling later, if 

needed. The effectiveness of this technique 

depends on the availability of reliable historical 

information on contamination, sources, bathymetry, currents, and other factors. 

2.2.X.3 Random 

The random sampling design is most useful 

for cases where little is known about the likely 

distribution of sediment contamination or sources, 

or when available information indicates a high 

degree of homogeneity in an area. The area to be 

sampled is divided using a grid system. Samples 

are distributed within the grid randomly, with each 

location having an equal probability of being 

sampled. The number of samples is selected 

statistically based on the requirements of the 

survey and the acceptability of false-positive or 

false-negative resuIts. This design yields statistically sound results. 

2.2.1.4 Stratijkd Random 

The stratified random design is a variation on 

the previous two designs. Available information 

is used to identify different zones that are likely to 

be similar in degree of contamination or other 

characteristics. Samples sites are then randomly 

selected within the different zones. This design 

also yields statistically reliable results. 

2.2.X.5 Exhaustive 

In the exhaustive design, an area is subdivided 

into equal-sized units, each of which is then 

sampled. This design yields a very complete 

characterization. However, this design is usually 

very costly because of the large number of samples that need to be collected. 

2.2.2 Numbers of Samples 

Statistics can be used to determine the number 

of samples .needed. To use statistics in this way, 

one needs to decide what comparisons will be 

niade with the resulting data and what will be the 

desired statistical power of the comparisons (i.e., 

at what level of confidence will resulting differences be tested). In addition, one needs some 

information about the inherent environmental 

variability in the area (i.e., the likelihood that an 

observed difference is due to an actual difference 

in contamination rather than just the natural 

heterogeneity in sediment or benthic population 

characteristics in the area). There are many 

different statistical approaches to estimating the 

number of samples required and to interpreting the 

resulting test results. Excellent reviews of statistical designs and interpretation are given by Baudo 

(1990) for sediment physical and chemical testing 

and by Downing and Rigler (1984) for benthic 

community structure evaluations. 

In practice, constraints on resources often 

preclude the use of a purely statistical approach to 

determining the number of samples and some 

form of a cost-benefit approach is often used to 

arrive at a reasonable compromise between statistical power and the cost of the study. One of the 

major advantages of the tiered approaches for 

testing and assessment is the cost savings that 

results when information is collected relatively 

inexpensively initially and additional resources are 

expended only when the information collected 

thus far is insufficient to make a decision. 

Guidance on how to select a co&effective 

approach is usually provided in very general 

qualitative terms as to the factors that should be 

considered in arriving at a decision (USEPA/ 

USACE, 1991; Higgins, 1988; Plumb, 1981). 

Decisions are largely subjective. However, researchers at EPA’s Environmental Research 

Laboratory (ERL)-Narragansett/Newport recently 

developed a four-step procedure to determine the 

optimal cost-effective sampling scheme for marine 

benthic community assessment (USEPA, undated). 

The procedure begins with an initial limited 

sampling using two or more sampling schemes at 

paired sites (test and reference sites). ‘Ihe “costs” 

in time and money are assessed for each sampling 

scheme. Next, a statistical power analysis is 

conducted to calculate the number of replicate 

samples needed to achieve a desired degree of 

statistical “power” for each sampling scheme. 

Finally, the power-cost efficiencies of the altemative sampling schemes are. calculated and the 

24 
2-QA/QC, Sampling, and Analytical Considetatbns 

optimum scheme is selected as the one with the 

highest power-cost efficiency. 

23 QUALITY ASSURANCE/QUALITY 

CONTROL 

Quality assurance and quality control (QA/ 

QC) are essential to the production of environmental monitoring data of known and documented 

quality in a cost-effective manner. QA/QC should 

be an integral part of the process of study design, 

execution, and data evaluation and interpretation. 

All EPA data-collection programs have implemented Quality Assurance Program Plans designed 

and overseen by their management to ensure the 

quality of all activities for which their organization is responsible. These programs address all 

quality assurance issues in regard to policy, 

planning, review, and implementation. QA Project Plans are a vital part of the QA Program Plan. 

A QA Project Plan is a project-specific guidance 

compiled to encompass all aspects of the sampling/analytical effort. The preparation of a QA 

Project Plan is often met with unnecessary trepidation. A QA Project Plan is simply a written 

record of the plans that must be made and followed in executing a study. A QA Project Plan 

provides detailed documentation of all facets of 

how and why a particular study will be undertaken. The Plan also describes the alternative actions 

that will be taken in the event that things do not 

go according to the original plans. Once all of the 

purposes and procedures of the proposed study are 

recorded in a QA Project Plan, the Plan can be 

improved or modified, if needed, through reviews 

by persons knowledgeable about different aspects 

of the study (e.g., chemical analysis, sampling 

logistics, navigational positioning, sample preservation techniques). 

Because the QA Project Plan is a vital tool for 

the datacollection process, it is essential that all 

personnel involved in the project read and understand the Plan and that the Plan be available for 

reference throughout lhe project to ensure proper 

implementation. 

QA Project Plans are important for legal as 

well as scientific reasons. QA Project Plans are 

required for all EPA-associated projeds (EPA 

Order 5360.1). QA Project Plans become part of 

contracts that are issued to undertake studies (40 

CFR, Part 15). Furthermore, nonadherence to the 

Plan could result in the data being unusable for 

court proceedings or regulatory decisions. 

The QA Project Plan is just as important after 

the study is completed and the data are being used 

to make an evaluation or decision. The Plan 

provides the information needed to assess the 

degree of confidence one can place in the data, as 

well as the comparability of the data collected in 

a particular study with those from another study. 

A common problem that managers and scientists 

have with using existing data is not that the old 

data are unreliable, but that the data are of unknown reliability. 

23.1 QA/QC Terminology 

A number of important concepts and terms 

need to be defined to develop an understanding of 

what makes up an adequate QA/QC program 

(USEPA, 1983; Delbert and Starks, 1985). 

Accumcy is defined as the difference between 

a measured value and the assumed or expected value. Accuracy in percent is 100 minus 

the total error, which is composed of bias and 

random errors. 

Bias is the systematic distortion of a measurement process that adversely affects the representativeness of the results. Bias can result 

from the basic sampling design, the kind of 

equipment used to collect the samples, the 

sample-handling procedures, and poor recovery of the analyte. Because bias is systematic, 

its magnitude can be predicted if proper QA 

procedures are being used in the field and 

laboratory. 

Comparczbility is the measure of confidence 

one has in being able to compare one data set 

with another. Comparability is increased if 

similar field and laboratory methods were 

used and decreased if different or unknown 

(undocumented) methods were used. Canparability between different laboratories can be 

evaluated through the use of inter-laboratory 

2-5 
Sediment Classification Methods Compendium 

comparisons, or “round-robin” studies, wherein standardized samples are analyzed by each 

of the participating laboratories. 

Completeness is the amount of valid data 

obtained (i.e., that met QAfQC acceptance 

criteria) compared to the planned amount. 

Completeness is usually expressed as a percentage. 

Du& quiz&y refers to the sum of all features 

and characteristics of the data that determine 

its capability to satisfy the objectives of the 

data collection. 

Du&a qua&y indicators are quantitative statistics and qualitative descriptors that are used to 

interpret the degree of acceptability or utility 

of data to the user. Data quality indicators 

include bias, precision, accuracy, comparability, completeness, and represenbtiveness. 

Data quality objectives (DQO) are statements 

of the overa uncertainty that a decisionmaker is willing to accept in results or decisions derived from the data, and they provide 

the framework for the data-collection effort. 

Duplicate samples are two samples taken 

from and representative of the same population and carried through all the same steps of 

sampling, storage, and analysis in an identical 

manner. 

Field blank is a clean sample (i.e., distilled 

water) carried to the sampling site, exposed to 

sampling conditions, and returned to the 

laboratory and treated as an environmental 

sample. Field blanks are used to try to assess 

contamination problems caused by conditions 

in the field, including contamination of the 

sampling device, sample containers, shipping 

containers, etc. 

Measurement error is the difference between 

the true sample values and the reported values 

and can occur during analysis, data entry, 

database manipulation, or other steps. 

Method sensitivitylmethod detection limit 

defines the lower limits of reliable analysis of 

a particular parameter inherent in the use of a 

particular test method. ‘Ihe method detection 

limit is the minimum concentratioh of a 

substance that can be measured with 99 percent confidence that the analyte concentration 

is flea&x than zero in a particular medium (40 

CFR Part 136, Appendix B). 

Precision is the degree of consistency among 

duplicate/replicate measurements. 

Qua& ussumnce is an integrated program 

for ensuring the reiiability of monitoring and 

measurement data. It includes the welldefined plans and procedures for how to ensure 

the production of sufficient data of known and 

documented quality, including monitoring how 

well QC procedures are actually being implemented. 

Quality control is the routine application of 

procedures for obtaining prescribed standards 

of performance in the monitoring and measurement process. It is the actual implementation of the QA plan, effected through measurements of data quality througb the use of 

blanks, spikes, etc. Quality control consists of 

both internal and external checks including 

repetitive measurements, internal test samples, 

interchange of technicians and equipment, use 

of independent methods to verify findings, 

exchange of samples and standards among 

laboratories, and use of standard reference 

materials. 

Run&m error is nonsystematic (and, therefore, unpredictable) error that can mr during any part of the sample collection, handling, and analysis. Hopefully, random errors 

are normally distributed with a mean of zero 

so that the overall evaluation will not be 

affected even though individual measurements 

will be affected. 

Representativeness is the degree to which the 

data accurately and precisely represent the 
2-QA/QC, Sampling, and Analytical Considerations 

parameter or condition being sampled. Representativeness is affected by sampling design 

(e.g., number of samples, method of selecting 

sampling sites), as well as analytical sampling 

accuracy and precision. 

Sampling error is the difference between the 

sampled value and the true value, and is a 

function of natural spatial and temporal variability and sampling design. It also includes 

error due to improperly selected/collected 

samples or improperly gathered measurements. 

Sampling error is more difficult to control 

than the other type of error, measurement 

error, and typically accounts for most of the 

total error. 

Uncertuinty is the total variability in sampling 

and analysis including systematic error (bias) 

and random error. 

Duplicates, spikes, and blanks are all used to 

assess the quality of the data, to identify any 

systematic problems, and to isolate the sources of 

such problems. 

2.4 SOURCES AND SIGNIFICANCE OF 

MONITORING ERROR 

To increase the accuracy, precision, and 

representativeness of the data collected in a sediment assessment study, it is important to be aware 

of and minimize two types of error that can be 

introduced into sediment contaminant concentration data: bias and scatter. Sources of bias in 

sediment studies include the actual heterogeneity 

in the distribution of contaminants in the sediments, the sampling design (number of samples, 

method for selecting sampling sites), the sampling 

method, the sample preparation procedures, and 

the testing methods. 

Factors that tend to make sediment contaminants distribute themselves heterogeneously 

include the differences in the density of the bulk 

contaminant (e.g., sinking versus floating); differences in the affinity of the contaminant for particles as a function of particle size, organic carbon 

content, etc.; particle sorting as a function of 

water currents and particle size; lateral mixing of 

water and sediments as a function of flow or 

distance downstream of the sources; resuspension; 

bioturbation; and biouptake. 

The objective of a well-designed sampling 

program is to minimize the introduction of data 

artifacts associated with the sampling plan, sample 

collection, sample preparation, and sample analysis while revealing the actual contaminant concentration profile in space as a function of time. A 

plan that requires preferential sampling of areas 

that are devoid of aquatic life will likely be biased 

toward high toxicant concentrations, resulting in 

an unrepresentative horizontal spatial sediment 

contaminant profile. Artifactual variability can be 

introduced if the number and size of the samples 

are inappropriate to the scale of the system under 

investigation, yet the sampling size has to be 

balanced against cost. 

With respect to bias due to sampling method, if 

certain core samplers are used to quantify the 

vertical distribution of a sediment contaminant, for 

example, the actual vertical profile is likely to be 

distorted because the absolute vertical relationship of 

contaminant concentrations is lost due to differential 

compression of the sample during coring. Another 

example of samphng method bias occurs when a 

grab sampler is used to collect the surficial sediment 

sample. The potential disproportionate loss of fme 

particles from the grab during the drop, closing, and 

withdrawal phases of sampling can result in an 

underquantification of the contaminant surticial 

concentration if the contaminant is preferentially 

concentrated on the fines. 

Regarding sample preparation bias, a sample 

preparation procedure that transforms, loses, or 

destroys one member of a homologous series (e.g., 

PCBs, PCDDs, or PCDFs) will not only result in 

an underquantification of the total concentration 

for that toxicant category, but will also misrepresent the relative proportions of the isomers. 

Analytical method bias can result from the inability to separate complex mixtures into individual 

constituents (interference), thus resulting in the 

misidentification or misquantification of a toxicant; from differences in the sensitivity of the 

2-7 
Sediment Classification Methods Compendium 

detector for a particular pollutant over the range of 

concentrations encountered in the sediment (nonlinear responses); or from poor or varying recovery of the analyte. 

Analytical variability arises primarily from the 

compounded uncertainty associated with the 

tolerance on each of the components and steps of 

the wet or electronic methods of sample preparation (aliquot selection, weighing, drying, grinding, 

sieving, etc.) and analysis. 

2.5 COMPONENTS OF A QUALITY 

ASSURANCE PROJECI’ PLAN 

As mentioned previously, a QA Project Plan 

clearly documents the participants’ responstbilities; 

what will be done; why it is being done; the 

desired accuracy, precision, completeness, and 

representativeness of the resulting data; who will 

report what information to whom; and what will 

be done in the event something goes wrong. 

Rather than attempting to describe the actual 

components of a QA Project Plan in any detail 

here, an example of the table of contents from a 

recent plan is presented in Figure 2-l. In addition, actual QA Project Plans from projects similar 

to the one being planned can be extremely useful 

in suggesting the important issues to consider. 

For detailed guidance on preparing QA Project 

Plans, one should refer to Interim Guidelines and 

Speci@ations for Preparing Quality Assurance 

Project Plans (USEPA, 1980). Some good 

examples of actual sediment assessment Quality 

Assurance Project Plans include Burton (1989), 

Ckcelius (1990), and Valente and Schoenherr 

(1991). 

2.6 SAMPLE COLLECTION AND 

HANDLING 

2.6.1 Sampling for Physical and Chemical 

Analyses 

2.6.1.X Sample Cdlection Meha 

The most appropriate device for a specific 

study depends on the study objectives, sampling 

conditions, parameters to be analyzed, and cost-effectiveness of the sampler. There are basically 

three types of devices used to collect sediment 

samples: dredges, grab samplers, and corers 

(Baudo, 1990). 

A died& is a vessel that is draggei3 across 

the bottom of the surface being sampled, coliecting a composite of surface sediments and associated benthic fauna. Dredge samplers are more 

commonly use4i to sample sediments in marine 

waters than in fresh water. This type of sampler 

is primarily used for collecting indigenous be&tic 

fauna rather than samples for analyses or assays. 

Because the sample is mixed with the overlying 

water, no pore water studies can be made of 

dredged samples. Additionally, because the walls 

of the dredge are typically nets, they ad as a sieve 

and only the coarser material is trapped, resulting 

in the loss of fme sediments and water-soluble 

compounds (ASTM, 1990). Results of dredge 

sampling are considered qualitative in nature since 

it is difficult to determine the actual surface 

sampled by the dredge. 

Grab samplers have jaws that close by a 

trigger mechanism upon impact with the bottom 

surface. Grab samplers offer the advantage of 

being able to collect a large amount of material in 

one sample, but they have the disadvantage of 

giving an unpredictable depth of penetration. 

Grab samplers are recommended when sampling 

is being performed for routine dredging projects 

because the sediments are continually disrupted by 

marine traffic, homogenizing the sediments that 

have accumulated since the last dredging (Plumb, 

1981). 

A core sampler is basically a tube that is 

inserted into the sediment by various means to 

obtain a cylinder or box sample of material at 

known depths. Corers can be simple, hand-operated devices used by scuba divers, or they can be 
2-QA/Qc, Sampling, and Analytical Cons&rations 

Cover Page (w/Approval Signatures) 

Tie Page 

Introduction 

Table of Contents 

List of Tables 

List of Figures 

List of Appendices 

List of Acronyms and Abbreviations 

GiOSStlQf 

1 PROJECT DESCRIPTION 

1 .l introduction 

1.2 Project Scope 

1.3 Data Quality Objectives 

1.4 Sample Network Design and Rationale 

1.5 Project Implementation 

2 PROJECT ORGANlZATlON AND 

RESPONSlBlLlT’f 

2.1 Organization 

2.2 Authority and Responsibility 

2.2.1 Project Oversight 

2.2.2 Field Activities 

2.2.3 Laboratory Analyses 

2.2.4 Other Regulatory Personnel 

2.3 Project Communication 

3 QUALITY ASSURANCE OBJECTIVES 

3.1 Quality Assurance Documents 

3.2 Project Quality Assurance Objectives 

3.3 Field Measurement Quality Objectives 

3.3.1 Navigation 

3.3.2 Sample Collection Parameters 

3.3.3 Water Column Measurements 

3.4 Laboratory Data Quality Objectives 

3.5 Macrobenthic Community Assessment 

Quality Assurance Objectives 

3.6 Computer Model Quality Assurance 

Objectives 

4 SAMPLE COLLECTION AND 

HANDLING PROCEDURES 

4.1 Sample Containers 

4.1 .l Volume and Type 

4.1.2 Quality Control and Storage 

4.2 Sampling Procedures 

4.2.1 Selection and Decontamination of 

Equipment 

4.2.2 Sampling Methods 

4.2.3 Collection of Sample 

4.2.4 Sample Volume, Presewation, and 

Holding Times 

4.2.5 Fold-Generated Waste Disposal 

4.3 Sample Packaging and Shipment 

5 SAMPLE DQCmENTATlON AND CUSTODY 

5.1 Field Procedures 

5.1.1 Sample Labeling 

5.1.2 Field Logbooks 

5.1.3 Field Chain of Custody 

5.1.4 Transfer of Custody 

5.2 Laboratory Procedures 

5.2.1 Sample Scheduling and Management 

5.2.3 Sample Receipt and Handling 

5.2.4 Log Books and Chain of Custody 

5.2.5 Sample Disposal 

5.3 Final Evidence File 

5.3.1 Contents 

5.3.2 Custody Procedure 

6 CAUBRATION PROCEDURES AND 

FREQUENCY 

6.1 Field Measurements 

6.1 .l Records and Traceability of Standards 

6.1.2 Initial and Continuing Calibration 

Procedures 

6.1.3 Conditions to Trigger Recalibration 

6.2 Physical and Chemical Laboratory 

Analyses of Sediment 

6.2.1 Records and Traceability of Standards 

6.2.2 Preparation and Storage of Standards 

6.2.3 Initial and Continuing Calibration 

.Procedure 

6.2.4 Conditions to Trigger Recalibration 

6.3 Biological Effects Tests -- Water Quality 

Monitoring 

6.3.1 Records and Traceability of Standards 

6.3.2 Initial and Continuing Calibration 

Procedure 

6.3.3 Conditions to Trigger Recalibration 

Figure 2-l. Contents of a Quality Assurance Project Plan 

2-9 
Sediment Chssification Methods Compendium 

7 MEASUREMENT PROCEDURE 

7.1 Field Measurements 

7.1.1 Navigation 

7.1.2 Sample Collection Parameters 

7.1.2.1 Sediment 

7.1.2.2 Fsh 

7.1.2.3 Benthic Organisms 

7.1.3 Water Column Measuremenk 

7.2 Chemical Analysis of Sediment 

7.2.1 Sample Preparation Methods 

7.2.2 Sample Extract Cleanup Methods 

7.2.3 Analytical Methods 

7.3 Other Sediment Analyses 

7.4 Biological Effects Tesk 

7.5 Macrobenthic Community Assessment 

7.6 Model Calculations 

8 INTERNAL QUALllY CONTROL CHECKS 

8.1 Sample Collection 

8.2 Field Measurements 

8.3 Chemical Analyses of Sediment 

8.4 Other Analyses of Sediment 

8.5 Biological Effects Tesk 

8.8 Macrobenthic Community Assessment 

8.7 Computer Model Calculations 

9 DATA REDUCTlON, VAUDATION, AND 

REPORTING 

9.1 Field Measurements 

9.2 Laboratory Data 

9.2.1 Internal Data Reduction 

9.2.2 Data Reporting Requirements 

9.2.3 External Data Validation 

9.3 Macrobenthic Community Assessment 

9.4 Computer Model Calculations 

10 PERFORMANCE AND SYSTEM AUDITS 

10.1 Audit Scheduling and Planning 

10.2 Internal Audits 

10.2.1 Field Activities 

10.2.2 Laboratory Activities 

10.2.2.1 System 

10.2.2.2 Performance 

10.3 External Audits 

10.3.1 Fold Acfivfttes 

10.3.2 Laboratory Activities 

10.3.2.1 System 

10.3.2.2 Performance 

10.4 Audit Reports 

11 PRNENTNE MAINTENANCE 

11.1 Fold Equipment 

11.2 Sample Collection Equipment 

11.3 Laboratory lnstrumenk 

11.4 Computer Hardware and Software 

12 SPEClFiC ROUTlNE PROCEDURES TO 

ASSESS DATA USABIUTY 

12.1 Sample Collection 

12.2 Field and Laboratory Data 

12.2.1 Data Duality Indicators 

12.2.1.1 Sensitivity 

12.2.1.2 Precision 

12.2.1.3 Accuracy 

12.2.1.4 Completeness 

12.2.2 Other Data Review 

12.3 Macrobenthic Community Assessment 

12.4 Computer Model Calculations 

13 CORRECTNE ACTlONS 

13.1 Introduction 

13.2 Equipment Failures 

13.3 Procedural Problems 

13.4 Sample Custody Failures 

13.5 Documentation Deficiencies 

13.6 Data Anomalies 

13.7 Performance Audit Failures 

13.8 System Audit Failures 

14 QUALllY ASSURANCE REPORTS TO 

MANAGEMENT 

14.1 Project-Specific Final Reports 

14.2 Deviation and Corrective Action Memos 

14.3 Internal and External Audit Reports 

15 REFERENCES 

Figure 2-l. Contents of a Quality Assurance Project Plan. (Continued) 
2caA/Qc, Sampling, and Analytical Considerations 

large, costly, motor-driven mechanisms that can 

collect samples from great depths. A few types of 

corers include a gravity corer, which uses weights 

attached to the head of the sampling tube to push 

the tube into the sediment; a piston corer, which 

is similar to a gravity corer but also has a piston 

inside the tube that remains stationary during sediment penetration and creates a vacuum that helps 

pull the sampler into the sediment; a vibra-corer, 

which is like a gravity corer except with a vibrating head attached to enhance penetration; and a 

multiple corer, which is an array of plastic tubes 

attached to a frame, allowing for the collection of 

several samples at the same location. Because 

gravity cOrers can compact the sample and distort 

the vertical profile, a piston corer or vibra-corer is 

recommended to minimize sample compaction. 

The corer that disturbs the sediments the least is 

a box corer. Instead of being cylindrical, it is a 

large box-shaped sampler that is deployed inside 

a frame. After the frame is brought to rest on the 

bottom, heavy weights lower the open-ended box 

into the sediment. A bottom door then swings 

shut upon retrieval to prevent sample loss. The 

advantages of the box corer include its ability to 

collect a large amount of sample with the center 

of the sample virtually undisturbed. Corers are 

not generally recommended for use in sandy sediments since they have difficulty retaining the 

sample upon withdrawal. 

A comparison of the general characteristics of 

various commonly used sediment-sampling devices for chemical, physical, and biological studies is 

given in Baudo (1990); Plumb (1981); Downing 

and Rigler (1984); and ASTM (1990). 

2.6.2 Sample Handling, Containers, 

Preservation, and Holding Times 

2.6.2. I General Requiremenf.s 

Proper handling of the samples is essential to 

preserve the sample integrity and the validity of 

the results. Mishandling of samples at any stage 

of the sample-collection process could distort 

analytical results, wasting the effort and expense 

of the sampling survey. Some of the basic considerations in sediment sample handling include 

the following (Plumb, 1981): 

n It is essential that noncontaminated sampling devices are used and that obvious 

sources of contamination such as exhaust 

fumes from the collecting ship, lubricating 

drilling fluids, and powder from surgical 

gloves be eliminated. 

m Sampling devices should be washed between samples with an appropriate series 

of cleansers and solvents to prevent crow 

contamination from one sample to the 

next. 

l Analysis for different parameters requires 

different storage containers to ensure 

noncontamination and to prevent degradation of the sample. Basic rules for 

containers include using plastic or glass 

containers for metal analysis, glass containers for organic analysis, and glass or 

plastic for inorganic analysis. Since no 

set guidelines have been determined for 

sediment sampling, a good general rule to 

follow is to use containers recommended 

for water testing. 

8 A reliable and identifiable sample-labeling 

process should be used. 

n Sampling containers should be filled to 

capacity, allowing only enough air space 

for possible expansion of the sample 

resulting from the preservation technique 

(e.g., freezing) to eliminate or greatly 

reduce oxidation of the sample (USEPA/- 

USACE, 1991). Sample containers for 

volatile organ& analyses should be filled 

completely, allowing no headspace. 

Preservation methods are intended to maintain 

the integrity of the sample by limiting the deter& 

ration or alteration of a specified parameter by 

hydrolysis, oxidation, and/or biological activity 

while the sample awaits analysis. Methods are 

basically limited to pH control, chemical addition 
Sediment CLrss$u&urr Methods Compntdium 

or fixation, sample extraction or isolation, or 

temperature control. Preservation steps should be 

initiated immediately after collection of the sample 

since significant alteration of the sample can occur 

in the first few hours after sampling. Immediately after collection, sediment samples are typically 

kept on ice or refrigerated. Upon arrival at the 

laboratory, samples are usually preserved by 

drying, freezing, or cold storage (ASTM, 1990). 

The type of preservation required will depend 

on the parameters being tested. For example, if 

the sediment is to be tested for both bulk metals 

and particle size, either two samples should be 

coliected or the sample should be split, since it is 

recommended that samples for bulk metal analysis 

be preserved by dry ice and stored at less than 

-2O”C, whereas samples to be analyzed for particle 

size should be refrigerated at 4°C (USEPA/ 

USACE, 1991). For this reason, it is essential to 

know which tests are to be performed, or potentially performed, on the samples in advance to 

allow for additional sample collection or splitting 

of samples as needed to comply with differing 

sampling, handling, and preservation requirements. 

Freezing appears to be the generally preferred 

method for preserving sediment samples for most 

chemical analysis, although sediments to be used 

for particle size determination, volatile organ@ 

and toxicity testing should not be frozen (ASTM, 

1990). 

2.6.2.2 Requirements for Specijic Analyses 

There are basically four ways to analyze 

chemical and physical parameters of sediments: 

bulk analysis, standard elutriate test, fractionation 

procedures, and physical analysis. Brief descriptions of these types of analyses follow, along with 

any special sample handling procedures, containers, or preservation techniques needed. 

Bulk anafysis aflows one to evaluate the total 

concentration of a parameter within a sediment 

sample or the toxicity of the whole sediment. 

Most chemical parameters are evaluated by bulk 

analysis. In general, the coIlection container and 

preservation and storage method are dependent on 

the parameter to be tested. Bulk analysis samples 

can be stored wet, air-dried, or frozen. If trace 

organic constituents are to be analyzed, a glass 

container should be used to store the sample. 

When preserving and storing samples, one needs 

to take into consideration that other parameters 

could change as a result of oxidation, volatilization, or chemical instability (Plumb, 1981). 

Eiutkte tes2s indicate the ability of chemical 

constituents to migrate from the solid phase to the 

liquid phase. An elutriate sample is Prepared by 

mixing or shaking sediment and water in prescribed proportions for a prescribed period of time 

and separating the liquid fraction by fdtration 

and/or centrifugation. The liquid fraction, the 

elutriate, is then analyzed by methods used for 

analysis of water samples. Sediments to undergo 

elutriate testing should be stored wet, at 4”C, in 

airtight containers and should be tested as soon as 

possible following sample collection. If trace 

organic analyses are to be performed, glass containers with Teflon lids are required for storage 

(Plumb, 1981). 

Fmctionation procedures provide information 

on the distrtbution of constituents. The samples 

are extracted multiple times using a series of 

extractants and procedures, thereby isolating 

specific pollutants or classes of pollutants. Pore 

water extraction is a form of fractionation whereby the interstitial water in the whole sediment 

sample is extracted by squeezing or centrifugation. 

The resulting water sample can be used in chemical and biological tests. To date, fractionation has 

been used primarily for research. As a result, 

most agencies do not subject their sediment 

samples to fractionation procedures (Plumb, 1981). 

However, some fractionation tests, such as the 

toxicity identification evaluation (TIE), a fractionation procedure to isolate the toxic component of 

a sample, are beginning to be used to make 

decisions regarding regulatory actions and remedial approaches since they can be used to assess 

which pollutants are responstble for the toxicity 

observed in a sediment. Samples to be analyzed 

for fractionation should be stored wet, at 4°C and 

in airtight containers. Testing procedures should 

start as soon as possible after sample colledion 

(Plumb, 1981). 

fhysicul anuiysis provides information on 

particle size, color, texture, and mineralogical 

2-12 
wA/QC, Sampling, and Analyticai Cons&ations 

characterization and includes tests for cation 

exchange capacity, particle size, pH, temperature, 

salinity, oxidation reduction potential, total volatile 

solids, and specific gravity. Samples to undergo 

physical analyses may be stored wet, at 4% or 

frozen, depending on the parameter to be tested. 

Some of these parameters (e.g., pH) should be 

analyzed immediately upor~ collection. 

The 1991 Green Book (USEPAAJSACE, 

1991) suggests the use of a grab sampler or corer 

for collection of sediment samples and offers the 

following general guidelines for preservation and 

handling and sample sizes needed for sediment 

samples collected for chemical and physical 

testing: 

Bulk metals should be stored in nonreactive 

containers, such as high-density polyethylene, and 

analyzed as soon as possible. 

Bulk orgaaics, including PCBs, pesticides, 

and high-molecular-weight hydrocarbons, should 

be contained in solvent-rinsed glass jars with 

Teflon lids, preserved by dry ice, and stored at 

less than -20°C in the dark. The samples can be 

stored for up to 10 days. Approximately 475 mL 

of sample should be collected. 

Samples to be analyzed for total organic 

carbon (TOC) should be preserved by dry ice and 

stored at less than -20°C. They can be kept for an 

undetermined amount of time. 

Sediments for particle size testing should be 

kept refrigerated at 4°C in any sealed container 

and can be kept for an undetermined amount of 

time. 

2.63 Minimum Parameters to Be Tested 

Sampling efforts are performed with a variety 

of objectives in mind, and therefore the minimum 

chemical and physical parameter testing requirements vary between studies or programs. However, some chemical and physical parameters seem 

to be common to several programs. They include 

particle or grain size, total organic carbon, heavy 

metals, acid volatile sulfides, polycyclic aromatic 

hydrocarbons, polychlorinated biphenyls, and 

pesticides. Unionized ammonia must also be 

measured, taking into account its sensitivity to pH 

and temperature, both of which are affected by 

sample manipulation. When testing sediment 

samples from estuarine or marine environments, 

the analysis methods chosen must address salinity 

since this can alter the analytical results (USEPA/- 

USACE, 1991). 

Particle or grain size analysis is a physical 

parameter that determines the distribution of 

particle sizes. Methods for particle size analysis 

are suggested in Folk (1%8), Buchanan (1984), 

Plumb (1981), ASTM (1990), and Te&a Tech 

(1985). Plumb (1981) suggests that analysis will 

usually require two or more methods, depending 

on the range of particle sizes encountered. He 

gives a detailed account of the use of sieves in 

conjunction with electronic particle counters or 

sieves and pipet analysis. Testing and Reporting 

Requirements for Ocean Disposal of Dredge 

Material ofl Southern CaZ~omiu under Marine 

Protection, Research and Sanctuaries Act Section 

103 Permits (Ocean Dredged Material Disposal 

Program, 1991) recommends the method given in 

Plumb (1981) for analysis of particle size. 

Total organic carbon (TOC) is an important 

indicator of bioavailability for nonionic hydrophw 

bit organic pollutants. When analyzing for this 

parameter, it is essential that the sample be stored 

in a glass or plastic container and that all air 

bubbles be removed from the sample before it is 

sealed and stored. The method given in Plumb 

(1981) is commonly recommended (Tetra Tech, 

1985). Plumb (1981) suggests using sample 

ignition, which uses a hydrochloric acid wash to 

separate the inorganic and organic carbon, or 

differential combustion, which uses thermal 

combustion to separate the two carbons by their 

different combustion temperature ranges. The 

1991 Green Book recommends that the analytical 

method to test for TOC be based on higb-temperature combustion rather than on chemical oxidation. 

Additionally, it recommends using sulfuric acid 

rather than hydrochloric acid rinse. Testing and 

Reporting Requirements for Ocean Disposal of 

Dredge Material off Sourhem California under 

Marine Protection, Research and Sanctuaries Act 

Section 103 Per?& recommends EPA Test 

Method No. 9060 for TOC determinations. The 

method recommended by EPA for use in applying organic carbon-normalized sediment quality 

2-13 
Sediment Classjtiurtion Methods Compendium 

criteria for nonionic hydrophobic organic chemicals uses catalytic combustion and nondispersive 

infrared detection (Leonard, 1991). 

Metak are found naturally occurring in the 

cavironrnent, but an excess of metals can be an 

indication of anthropogenic contamination. The 

most commonly used method to analyze sediments 

for metals is atomic absorption spectrophotometry. 

Plumb (1981) details the use of the direct-flame 

atomic absorption method for all metals except 

arsenic, mercury, and selenium. For these metals, 

he recommends using arsine generation, cold 

vapor technique, and digestion/flameless atomic 

absorption or hydride generation, respectively. 

The 1991 Green Book points out that the concentration of salt in marine or estuarine samples may 

cause interference in analysis for metals. Therefore, the approach of an acid digestion followed 

by atomic absorption spectroscopy should be 

coupled with an appropriate technique to control 

this interference. The 1991 Green Book recommends USEPA (1986) for analysis of mercury and 

EPRI (1986) for (he analysis of selenium and 

arsenic. Testing and Reporting Requirements for 

Ocean Disposal of Dredge Material ofi Southern 

California under Marine Protection, Research and 

Sanctuaries Act Section 103 Permits recommends 

the following EPA Test Methods: cadmium (Nos. 

7130, 7131); hexavalent chromium (Nos. 7190, 

7191); copper (No. 7210); lead (Nos. 7420,742l); 

mercury (No. 7471); nickel (No. 7520); selenium 

(Nos. 7740,7741); silver (No. 7760); and zinc 

(No. 7950). 

Acid volatile sulfides (AVS) have been found 

to be closely related to the toxicity of sedimentassociated metals (Di Toro et al., 1990). AVS 

have been found to be important in binding 

potentially bioavailable metals, thereby reducing 

their toxicity. The approved method is given in 

USEPA (1991). 

Polyaromatic hydrocarbons (PAHs) are 

semivolatile organic priority pollutants, a number 

of which are potential carcinogens. Plumb (1981) 

details the methods of methanol extraction/W 

analysis and ethanol extraction/UV spectrophotomttry to analyze for this parameter. Testing und 

Rep&g Requirements for Ocean Disposal of 

Dredge Material off Southern California under 

Marine Protection, Research and Sanctuaries Act 

Section 103 Permits recommends EPA Test 

Method Nos. 8100,825O and 8270 for analysis of 

PAHs. 

Polycblorinated blphenyk (FCB) are chlorinated organic compounds that were once used for 

numerous purposes including as a dielectric fluid 

in electrical transformers. Desirable properties of 

PCBs include low flammability, nonconductivity, 

and nonreactivity. However, PCBs do not break 

down readily and they biticcumulate in the 

environment. The 1991 Green Book offers gas 

duomatography/electron-capture detection (GC/ 

ECD) methods as the primary tool for the analysis 

of PCBs, or the use of GC/MS using selected ion 

monitoring (SIM). They do not recommend the 

traditional methods of PCB analysis, which quantify PCBs as arochlor mixtures. Testing and 

Reporting Requirements for Ocean Disposal of 

Dredge Material ofl Southern Califontia under 

Marine Protection, Research and Sanctuaries act 

Section 103 Permits recommends the use of the 

methods described in Tetra Tech (1986) and 

NOAA (1989) for analysis of PCBs. 

Pesticides are man-made compounds predominantly used in agriculture to control crop 

damaging insects. Some pesticides, especially 

halogenated compounds, persist in the environment and can contaminate the food chain. Plumb 

(1981) details the method of bexane extraction in 

preparation for testing for organophospborus 

pesticides. The 1991 Green Book recommends 

using GC/ECD or GC/MS to analyze for chiorinated pesticides. Testing and Reporting Requirements for Ocean Disposal of Dredge Material off 

Southern California under Marine Protect& 

Research and Sanctuaries Act Section IO3 Permits 

recommends EPA Test Method No. 8080 to 

analyze for pesticides. 

For analyses of volatile organic pollutants 

and semivolatile organic pollutants, the 1991 

Green Book recommends the methods descrii 

‘by Tetra Tech (1986), which should always 

include the use of capillary-column GC or GC&fS 

techniques. For volatiles, a purge-and-trap method 

is used, followed by GUMS analysis accmiing to 

U.S. EPA Method 624 or U.S. EPA Method 1624, 

Rev. B, Ref. 3 (Tetra Tech, 1986). 

2-14 
2-QA/QC, Sumpiing, and Analytical Considerations 

As stated previously, the minimum set of 

parameters tested in sediments varies and is based 

on the sampling objectives of the program. Listed 

below are several examples of minimum data sets 

required by specific programs. 

The 1991 Green Book recommends that all 

sediment samples be analyzed for TOC, PAHs, 

grain size, total solids/water content, and specific 

gravity. The remaining parameters to be sampled 

are compiled from the priority pollutants list based 

on historical testing data, potential contaminants 

due to known industries in the area, and a general 

knowledge of the area to be sampled. 

Testing and Reporting Requirements for 

Ocean Disposal of Dredge Material oflSourhem 

California under Marine Protection, Research and 

Sanctuaries Act Section 103 Permits has very 

specific parameters and methods required for 

materials to be disposed of off the coast. Required analyses for physical parameters include 

grain size, total solidWater content, and specific 

gravity. Chemical analyses includes 9 metals, 

ammonia, arsenic, total sulfides, acid volatized 

sulfides (AVS), 11 pesticides including total pesticides, 9 organic compounds, all PCB congeners, 

individual totals of tetra-, penta- and hexa-chlorobiphenyl isomers, and 17 PAHs. 

The EPA Environmental Monitoring and 

Assessment Program -Near Coastal (EMAP-NC) 

established guidelines identified in itsNear Coastal Program Plan for 1990: Esluaries (Holland, 

1990) for sediment sampling for determination of 

contaminant levels. They include sample collection by means of a Young-modified Van Veen 

grab and, initially, analyzing the NOAA Status 

and Trends suite of contaminants, which include 

chlorinated pesticides, PCBs, PAHs, major elements, and toxic metals. EMAP-NC, with the 

assistance of other programs, plans to refine the 

list of contaminants to include pesticides and 

herbicides and other toxic chemicals. 

2.6.4 Sampling for Benthic Community 

Structure in Fresh Water 

Macrobenthic organisms play an important 

role in marine, estuarine, and freshwater lotic and 

lentic ecosystems. As major secondary consumers, they represent an important linkage 

between primary producers and higher trophic 

levels for both planktonic and detritus-based food 

webs. They are a significant food source for 

juvenile fish and crustaceans and may improve 

water quality by filter-feeding of particulate matter 

(Holland, 1990). Benthic populations also represent diverse taxa and can seme as sentinels for 

environmental stress. Benthic organisms access 

all aspects of the aquatic habitat with varying 

feeding strategies, reproductive modes, life history 

charaderisks, and physiological tolerances to 

environmental conditions. Most benthic organisms have limited mobility and cannot avoid 

environmental stressors. As a result, the responses 

of some species serve as indicators of changes in 

sediment quality (Holland, 1990). This section 

will detail specific procedures and precautions 

necessary for proper conduct of benthic sample 

collection and handling in freshwater, marine, and 

estuarine ecosystems. 

2.6.4.1 Sample Collection Metho& 

It is helpful to consult Macroinvertebrati 

Field and Laboratory Methods for Evaluating the 

Biological Integrity ofsurface Waters (Klemm et. 

aL, 1990), which thoroughly addresses methodologY. State environmental regulatory programs 

should have a Quality Assurance Program Plan 

describing the field methods and standard operating procedures for collecting and evaluating 

benthic maaoinvertebrates. This information 

should be obtained to ensure acceptance and 

comparability of study results with those obtained 

by the state agency. If this information is not 

available, then field methods and standard operating procedures from other existing programs 

should be used. 

In soft freshwater sediments, the most common method used to collect benthos is with a grab 

sampler such as a Ponar (15 x 15 an or 23 x 23 

an) or Ekman grab sampler (15 x 15 cm, 23 x 23 

an, or 30 x 30 cm), each of which provides a 

quantitative sample based on the surface area of 

the sampler. The smaller of the sampler sizes are 

most’ commonly used for freshwater studies 

be-cause of their relative ease of manipulation. 
Se&rent Classification Methods Compendium 

The Ekman grab sampler is not as effective in 

areas of vegetative debris but is much lighter than 

the Ponar and easier to use in softer substrates. 

Artificial substrates (Hester-Dendy using several 

3-inch plates and spacers attached by an eyebolt, 

or substrate/rock-filled baskets) provide consistent 

habitat for the benthos to colonize in both softbottomed and stony areas. Artificial substrates 

can be used in almost any water body and have 

been successfully used to standardize results 

despite habitat differences (Ohio EPA, 1989; 

Rosenberg and Resh, 1982; and Resh and Jackson, 

1991). 

A variety of methods for sampling benthos in 

hard-bottomed lotic systems are available, including artificial substrates. If quantification by 

sediment or sampler surface area is needed, a 

Surber-type square-foot sampler with a Standard 

#30-mesh (0.589~mm openings) can be used 

(Klemm et al., 1990). The traveling kick-net (or 

dip-net) method, also using a #30-mesh net, can 

be used to quantify the sample collected by the 

amount of time spent sampling and the approximate surface area sampled (Pollard, 1981; Pollard 

and Kinney, 1979). The Surber-type and kick 

methods can each be used to provide consistent, 

reproducible samples, but both are limited to 

wadable streams. The Surber sampler’s optimal 

effectiveness is limited to riffles, whereas kick-net 

or dip-net samplers can be effectively used in all 

available habitats. Although dip-net samplers 

have been effectively used to sample riffles and 

other relatively shallow habitats to determine taxa 

richness, presence of indicator organisms, relative 

abundances, similarity between sites, and other 

information, they do not provide definitive estimates of the number of individuals or biomass per 

surface area. 

2.6.4.2 Sample Handling and Preservation 

The following decisions will need to be made 

once the sample collection method is chosen: 

(1) whether samples will be picked from debris 

and sorted in the field, (2) which preservative 

should be used, (3) whether a stain (rose bengal) 

or other material will be added to the sample to 

facilitate separating the organisms from debris, 

(4) the type of sample containers and labeling of 

the containers required, and (s) the mode of 

transportation of the samples to their destination. 

Many of these decisions are based on professional 

preference or the required logistics of the study. 

Sorting of the benthos from debris and preservation are fully discussed by Klemm ef al. (1990). 

American Public Health Association et al. (1989) 

and Klemm r~ al. (1990) defmed the benthos by 

what is retained cm a standard #30 sieve. However, some types of Chironomidae and other small 

benthos pass through a #30-mesh sieve but are 

retained by a #NJ-mesh sieve. It has been recommended that samples should fist be passed 

through a #3CLmesh sieve. Then the materials 

washed through should be passed through a #4Omesh sieve, and the materials retained in both 

sieves should be sorted (Ohio EPA, 1989). Once 

the material is washed through the sieves, the 

organisms should be separated from the vegetation 

and other debris in a white enamel pan. As the 

materials are separated, the organisms can be 

placed in different vials for the major taxa. 

Preservation with either formalin or 70 percent 

ethanol is common. Although formalin is an 

excellent fmative, the human health concerns 

associated with its use require extreme caution and 

adequate ventilation. Many programs rely on 70 

percent ethanol as a fixative and preservative. 

A practical technical reference that. details 

procedures for cost-effective biological assessments of lotic systems has been developed. Rapid 

Bioassessment Protocols for Use in Streams and 

Rivers: Benthic Macroinwrtebrates and Firh 

(Plafkin et al., 1989) presents three benthic rapid 

bioassessment protocois (RBPs) and two fish 

RBPs, with a progressive order of increasing rigor 

in evaluation within each series for each class of 

organisms. 

The RBPs are based on integrated assessments 

that compare physical conditions of habitat (e.g., 

physical structure, flow regime) and biolo@a.l 

measures of reference conditions. These reference 

conditions are derived after systematic monitoring 

of sites that represent the natural range of var& 

tion. in water chemistry, habitat, and biological 

condition. 

2-16 
Z-QA/QC, Sampling, and Analytical Cons&rations 

The functional and structural components 

evaluated for aquatic communities comprise eight 

metrics for benthic RBPs and 12 metrics for the 

fish RBPs. Examples of metrics for benthic 

communities include the following: taxa richness, 

the modified Helsenhoff Biotic Index (summarizes 

overall pollution tolerance of the benthic arthropod 

community with a single ,value; this index was 

modified to include nonarthropod species as well), 

ratio of scraper and filtering collector functional 

feeding groups, ratios of the number of organisms 

in the EPT (Ephemeroptera, Plecoptera, and 

Trichoptera) to the number of Chironomidae 

present, and community similarity indexes. The 

fish protocol is based on the index of biotic 

integrity (IBI) or a fish community assessment 

approach developed by Karr et al. (1981). As 

with the approach of metrics in the benthic evaluations, the metrics of the fish protocol represent 

differing sensitivities. 

2.6.5 Sampling for Benthic Community 

Structure in Marine and Estuarine 

Waters 

Historically, regional monitoring programs 

have used benthic community studies as an effective indicator of the extent of pollution impacts on 

marine and estuarine ecosystems, as well as the 

effectiveness of management actions. In addition, 

information on changes in benthic population and 

community parameters due to sediment characteristics can be used to distinguish natural variation 

from changes due to human activities (Holland, 

1990). 

2.6.5.1 Sample Colkction Methods 

Three grab samples are collected for benthic 

species composition, abundance, and biomass. 

Additional sediment grabs are collected for chemical analyses and for use in acute toxicity tests. To 

minimize the possibility of biasing results, benthic 

biology grabs should not be coIlected consecutively, but rather interspersed among the chemistry/toxicity grabs. While a biology grab is being 

processed (sieved), grabs should be colleckd for 

chemistry/toxicity (Holland, 1990). 

A ‘& m’, stainless steel, Young-modified Van 

Veen grab sampler may be used to collect sediments for benthic analyses. The sampler is constructed entirely of stainless steel and has been 

coated with Kynar (similar to Teflon) and is, 

therefore, appropriate for collecting sediment 

samples for both biological and chemical analyses. 

lhe top of the sampler is hinged to allow for the 

removal of the top layer of sediment for chemical 

and toxicity analyses. This gear is relatively easy 

to operate and requires little specialized training. 

To minimize the chance of sampling the exact 

same location lwice, the boat should be moved 

5 meters downstream after three grabs have been 

taken, whether successful or not (Holland, 1990). 

2.6.5.2 Sample Handling and Preservation 

Grab samples to be used in the assessment of 

macroinvertebrate communities are processed by 

first extracting a core sample from the sampler. 

The depth of sediment at the middle of the sampler should be at least 7 an. Descriptive information on the grab is recorded. The depth to the 

black layer of sediment within the core, the redox 

potential discontinuity (RPD), is measured in the 

field. The sample is then extruded from the core 

tube to fill a whirl pat bag, labeled, and recorded. 

The sample should be refrigerated at 4”C, not 

frozen (Holland, 1990). 

The remainder of the grab is processed for 

benthic community analysis. The sediments are 

transferred into a basin and then into a O&run 

mesh sieve. The sieve is agitated to wash away 

sediments and leave organisms, detritus, sand 

particles, and pebbles larger than 0.5 mm. A 

gentle flow of water over the sample is acceptable, 

but forceful jets of water should be avoided 

because they can cause mechanical damage to 

fauna. The organisms are rinsed and transferred 

from the sieve into a jar and covered in seawater 

with MgCl added. This “relaxes” the organisms, 

reducing damage from addition of the preservative 

(Holland, 1990). Ten percent buffered formalin is 

used to fii and preserve samples. After 30 minutes in the relaxant, formalin with a small amount 

of borax should be added to each sample jar. The 

jar is filled to the rim with seawater to eliminate 

2-17 
Scdimmf Classification Me&n& Compendium 

any air space, eliminating the problem of organisms sticking to the cap during shipment. Prior to 

sieving the next sample, the sieve is rinsed and 

brushed thoroughly to prevent cross-contamination 

of samples. 

2.6.6 Sampling for Bioassays and Toxidty 

TeSting 

Environmental impacts on marine ecosystems 

are primarily assessed and monitored using the 

tools outlined in the 1991 Green Book. The 1991 

Green Book is used to make decisions regarding 

the suitability of dredged material for ocean 

dumping. EPA and the USACE have shown that 

the greatest potential for environmental impact 

from dredged material disposal is on the benthic 

environment since benthic organisms burrow into 

and are exposed to sediments and associated 

contaminants for extended periods of time. The 

1991 Green Book uses whole sediment bioassays 

to evaluate potential impacts of dredged sediments 

and, in concert with the identification of contaminants of concern through chemical analysis, serves 

to determine the extent and type of bioavailability. 

In addition, sediment toxicity tests can be used to 

assess spatial and temporal changes in toxicity in 

contaminated areas, rank sediments based on their 

toxicity to benthic organisms, and define cleanup 

goals for contaminated areas. This section will 

highlight some of the collection and handling 

methods of sediments for toxicity testing and 

whole sediment bioassays. 

2.6.6.1 Sample Collection, Handling, and 

Preservation 

The sediment environment is composed of 

many microenvironments, redox gradients, and 

interacting physicochemical and biological processes. Many of these characteristics influence 

sediment toxicity and bioavailability to benthic 

and planktonic organisms, microbial degradation, 

and chemical sorption. Maintaining the integrity 

of a sediment sample during its removal, transport, 

storage, and testing in the laboratory is extremely 

difficult. Any disruption of this environment 

complicates interpretations of treatment effects, 

causative factors, and in situ comparisons (ASTM, 

1990). 

Sample handling, prmation, and storage 

techniques have to be designed to minimize any 

changes in composition of the sample bJ retarding 

chemical and/or biological activity and by avoiding contamination. Sufficient sample volume 

must be collected to perfom~ the necessary analyses, partition the samples for respective storage 

requirements, and archive portions of the sample 

for possible later analysis. Core sampling is 

recommended to best maintain the integrity of the 

sediment for studies of sediment toxicity, interstitial waters, microbiological processes, and 

chemical fate. Subsampling, cornpositing, or 

homogenization of sediment samples may be 

necessary depending on the study objectives. 

Subsamples of the inner core area may be taken 

since this area is more likely to retain its integrity 

and depth profile and not be contaminated by the 

sampler. The loss of sediment integrity and depth 

profile is an important consideration, as are cbanges in chemical speciation through oxidation and 

reduction resulting in volatilization, sorption, or 

desorption; changes in biological activity; completeness of mixing; and sampling container 

wntamination (ASTM, 1990). 

Subsamples of the top 1 or 2 an may be 

collected with a nonreactive sampling tool (e.g., 

polytetrafluoroethylene (PTF)-lined calibration 

scoop). Some studies may require a composite of 

single sediment samples, which usually consist of 

three to five grab samples. Subsamples should be 

collected with a Teflon paddle, placed in a nonreactive bowl or pan, and stirred until the texture 

and color appear uniform. The sediments should 

be removed and partitioned for chemical and AVS 

analysis. Samples should completely fill the 

storage containers, leaving no airspace. If the 

sample is to be frozen, just enough air space 

should be allowed for expansion to take place. 

The labeling system should be tested prior to use 

in the field, making sure that labels can withstand 

soaking, drying, and freezing without becoming 

detached or illegible (USEPA/USACE, 1991). 

Maintaining clean and uncontaminated sampling equipment between samples is necessary. It 

is important to clean the sampling device, scoop, 

2-D 
2caA/Qc, Sampling, and Analytical Considerations 

spatula, and/or mixing bowls between sites. A 

suggested cleaning procedure includes a soap-andwater wash followed by an organic solvent rinse 

(ASTM, 1990). 

The choice of sample containers for sediment 

should consider the type of sediment, storage time, 

chemical sorption, and sample composition. For 

sediments containing organics, brown borosilicate 

glass containers with Teflon lid liners are optimal, 

whereas plastic or polycarbonate containers are 

recommended for metal-containing sediments. 

PTF or high-density polyethylene containers are 

relatively inert and are suggested for use with 

samples contaminated with multiple chemical 

types (ASTM, 1990). 

Sediment samples for biological testing should 

be press-sieved through a l-mm mesh screen to 

remove all living organisms from the sediment 

prior to testing. Other matter retained on the 

screen with the organisms, such as shell fragments, gravel, and debris, should be recorded and 

discarded. Sediment samples for use in bioassays 

should be well mixed. 

Since the first few hours are the most critical 

to changes in the sample, preservation steps 

should be taken immediately upon sediment 

collection. There is no universal preservation or 

storage technique, and a technique for one group 

of analyses may interfere with other analyses. 

Problems can be overcome by collecting sufficient 

sample volume to use specific preservation or 

storage techniques for specific analytes or tests on 

subsamples. Preservation, whether by refrigeration, freezing, or addition of chemicals, should 

be accomplished in the field whenever possible. 

If final preservation techniques cannot be implemented in the field, samples should be temporarily 

preserved in a manner that retains the integrity of 

the sample. Sediment samples for biological 

analysis should be preserved at 4”C, never frozen 

or dried. Field refrigeration is easily accomplished with coolers and ice; however, samples 

should be segregated from melting ice or cooling 

water. 

Storage containers can be the same as the 

transport containers, and where sediments contain 

volatile compounds, transport and storage should 

be in airtight PTF or glass containers with PTFlined screw caps. Exposure of sediments to air 

should also be prevented in the handling of AVS-- 

containing sediments. AVS is the reactive sulfide 

pool that can reduce metal toxicity by binding 

metals in anoxic sediments. Oxidation of these 

sediments can either increase toxicity by disassociation of the AVS-metal complex and precipitation 

of the metal species, or reduce toxicity if the 

AVS-metal complex should volatilize (ASTM, 

1990). 

It has been found that sediments can be stored 

at ‘4°C without significant alterations in toxicity. 

Completion of testing within a 2-week storage 

period is recommended, but limits on storage 

time will depend on sediment and contaminant 

characteristics (ASTM, 1990). 

2.7 REFERENCES 

American Public Health Association, American 

Water Works Association, and the Water 

Pollution Control Federation. 1989. Standard 

methods for the examination of water and 

wastewater. 17th edition. APHA, Washington, 

DC. 

ASTM. 1990. Standard guide for colledion, 

storage, characterization, and manipulation of 

sediments for to5cicological testing. American 

Society of Testing and Materials. A!5I’M 

Designation E 1391-90. 

Baudo, R. 1990. Sediment sampling, mapping, 

and data analysis. pp. 15-60. In: R. Baudo, 

J.P. Giesy, H. Muntau, Sediments: Chemistry and Toxicity of In-Place Pollutants. 

Buchanan, J.B. 1984. Sediment analysis. In: 

Methods for the Study of Marine/Benthos. 

IBP Handbook No. 16, 2nd edition. NA 

Holme and A.D. McIntyre (eds.). Blackwell 

Scientific Publications, Oxford, UK. 

Buxton, G.A., Jr. 1989. Quality assurance projed 

plan for “A multi-assay/multi-test site evaluation of sediment toxicity.” U.S. Environmental Protection Agency, Great Lakes Nationat 

Program Office. 

Crecelius, E. 1990. Quality assurance projed 

plan for “Assessment and remediation of 

contaminated sediments (ARCS) assistance.” 

2-19 
Srdirnenf Classification Methods Compendium 

Prepared by Battelle/Marine Sciences Iaboratory for the Environmental Protection Agency, 

Great Lakes National Program Office. 

Delbert, SB, and T.H. Starks. 1985. Projed 

summary sediment sampling quality assurance 

user’s guide. EPA-600/4-85048. Prepared 

by Environmental Research Center, University 

of Nevada, Las Vegas, for U.S. Environmental Protection Agency, Office of Research and 

Development, Environmental Monitoring 

Systems Laboratory, Las Vegas, Nevada, May 

1985. 

Di Toro, D.M., J.D. Mahony, D J. Hansen, KJ. 

Scott, W. Burry, M.B. Hicks, SM. Mayr, and 

M.S. Redmond. 1990. Toxicity of cadmium 

in sediments: The role of acid volatile sulfides. In: Environmental Toxicology and 

Chemistry. In press. 

Downing, JA. and F.H. Rigler, eds. 1984. A 

manual on methods for the assessment of 

secondary productivity in fresh waters. 

Second edition Blackwell Scientific Publications. 

EPRI. 1986. Speciation of selenium and arsenic 

in natural waters and sediments. Vol. 2. 

Prepared by Battelle Pacific Northwest Laboratories for the Electrical Power Research 

Institute. EPRI EA-4641. 

Folk, R.L. 1968. Petrology of sedimentary rocks. 

University of Texas, Austin, TX. 

Higgins, T.R. 1988. Techniques for reducing the 

costs of sediment evaluation. Tech. Note 

EEDP-06-2. U.S. Army Engineer Waterways 

Experiment Station, Vicksburg, Mississippi. 

Holland, A.F., ed. 1990. Environmental monitoring and assessment program, near coastal 

program plan for 1990: Estuaries. Environmental Research Laboratory, U.S. Environmental Protection Agency. 

UC. 1988. Procedures for the assessment of 

contaminated sediment problems in the great 

lakes. Report lo the Water Quality Board of 

the International Joint Commission by the 

Sediment Subcommittee and its Assessment 

Work Group, International Joint Commission, 

Windsor, Ontario Canada, December, 1988. 

Karr, J.R., K.D. Fausch, P.L. Angermeier, P.R. 

Yant, and I.J. Schlosser. 1986. Assessing 

biological integrity in running waters: A 

method and its rationale. Illinois Natural 

History Survey, Special Publication 5. Springfield, IL 

Klemm, DJ., PA. Lewis, F. Fiulk, ‘and J.M. 

Lazorchak. 1990. Macroinvertebrate Iieldand 

laboratory methods for evaluating the biological integrity of surface waters. U.S. Environmental Protection Agency, Office of Research 

and Development, EPA/600/4-9OKl30. 

Lee, H. II, B.L. Boese, J. Pellitier, M. Winsor, 

D.T. Specht, and,R.C. Randall. 1989. Guidance manual: Bedded sediment bioaccumulation tests. U.S. Environmental Rote&on 

Agency, Pacific Ecosystems Branch, Bioaccumulation Team, Newport, Oregon. EPA-mO/x-89-302. ERLN-Nlll. 

Leonard, E. 1991. Standard operating procedures 

for total organic carbon analysis of sediment 

samples, U.S. Environmental Protection 

Agency, Office of Research and Development, 

Environmental Research Laboratory, Duluth, 

Minnesota. 

NOAA. 1989. Standard analytical procedure of 

the noaa national analytical facility. 2nd ed. 

NOAA Tech. Mem. NMFC F/NWC-92, 

19851986. 

Ocean Dredged Material Disposal Program. 1991. 

Testing and reporting requirements for ocean 

disposal of dredged material off southern 

califomia under marine protection, research 

and sanctuaries a& section 103 permits. 

Ohio EPA. 1989. Biological criteria for the 

protection of aquatic life: Volume III. Standardized biological field sampling and laboratory methods for assessing fish and macroinvertebrate communities. Division of Water 

Quality Planning and Assessment, Ecological 

Assessment Section, Columbus, Ohio. 

Plafkin, J.L, M.T. Barbour, KD. Porter, SK 

Gross, and R.M. Hughs. 1989. Rapid bioassessment protocols for use in streams and 

rivers: Benthic macroinvertebrates and f&. 

U.S. Environmental Protection Agency, Offke 

of Water, EPA/444(440)/4-39-001, Washington, DC. 

Plumb, R.H., Jr. 1981. procedure for handling 

and chemical analysis of sediment and water 

2-20 
2-QA/fjC, Sampling, and Analytical Considerations 

samples. Tech. Rep. EPA/CEdl-1. Prepared by Great Lakes Laboratory, State University College at Buffalo, NY, for the U.S. 

Environmental Protection Agency/Corps of 

Engineers Technical Committee on CXteria 

for Dredged and Fill Material. Published by 

the U.S. Army Engineer Waterways Experiment Station, Vicksburg, Mississippi. 

Pollard, J.E. 1981. Investigator differences 

associated with a kicking method for sampling 

macroinvertebrates. J. Freshwater Ecol. 

1:215-224. 

Pollard, J.E., and W.L. Kinney. 1979. Assessment of macroinvertebrate monitoring techniques in an energy development area: A test 

of the efficiency of three macroinvertebrate 

sampling methods in the White River. U.S. 

Environmental Protection Agency, Office of 

Research and Development, Las Vegas, NV. 

EPA-6(X)/7-79/163. 

Resh, V.H., and J.K. Jackson. 1991. Rapid 

assessment approaches to biomonitoring using 

benthic macroinvertebrates. In: Freshwater 

Biomonitoring and Benthic Macroinvertebrates. D.M. Rosenberg and V.H. Resh (eds.). 

Chapman and Hall, New York Press. 

Rosenberg, D.M., and V.H. Resh. 1982. The use 

of artificial substrates in the study of freshwater benthic macroinvertebrates. In: Artificial 

Substrates. J. Cairns, Jr. (ed.). Ann Arbor 

Science Publisher, Ann Arbor, MI. 

Ryti, R.T., and D. Neptune. 1991. Planning 

issues for superfund site remediation. Hazardous Material Control, November/December, 1991. pp. 47-53. 

Tetra Tech. 1985. Summary of U.S. EPA-approved methods, standard methods, and other 

guidance for 301(h) monitoring variables. 

Final Report, EPA Contract No. 68-01-6938. 

Tetra Tech. 1986. Analytical methods for U.S. 

EPA priority pollutants and 301(h) pesticides in estuarine and marine sediments. 

Final Report, EPA Contract No. 68-01-69- 

38. 

USEPA/USACE. 1991. Evaluation of dredged 

material proposed for ocean disposal-testing 

manual. U.S. Environmental Protection 

Agency and U.S. Army Corps of Engineers. 

USEPA. 1980. Interim guidelines and specifications for preparing quality assurance 

project plans, U.S. Environmental Protection 

Agency, Office of Monitoring Systems and 

Quality Assurance, Office of Research and 

Development, Publication Number QAMS005/80, December 29, 1980. 

USEPA. 1983. Guidelines and specifications for 

preparing quality assurance program plans. 

U.S. Environmental Protection Agency, 

Office of Research and Development, Quality 

Assurance Management Staff. 

USEPA. 1986. Test methods for evaluating solid 

waste. U.S. Environmental Rotection Agency, Office of Solid Waste and Emergency 

Response, Washington, DC. 

USEPA. 1991. Draft analytical method for 

determination of acid volatile sulfide (AVS) in 

sediment, proposed technical basis for establishing sediment quality criteria for nonionic 

organic chemicals using equilibrium partitioniqg, U.S. Environmental Protection Agency, 

Criteria and Standards Division, Washington, 

DC. 

USEPA. undated. Cost-Efficient sampling 

schemes for marine benthic communities. 

U.S. Environmental Protection Agency, 

Environmental Research Laboratory - Narragansett and Environmental Research Laboratory 

Newport, Publication Number ERLN-NlS6. 

Valente, R., and J. Scboenherr. 1991. Environmental monitoring and assessment program, 

near coastal Virginian Province, quality asurante project plan. Environmental Research 

Laboratory, U.S. Environmental Protection 

Agency. 

2-21 
CHAPTER 3 

Bulk Sediment Toxicity Test Approach 

Nelson Thomas 

U.S. Environmental Protection Agency, Environmental Research laboratory 

6201 Congdon Blvd., Duluth, MN 55804, (218) 720-6702 

Janet O. Lamberson and Richard C. Swartz 

U.S. Environmental Protection Agency, Pacific Ecosystems Branch, ERL-N 

2111 SE Marine Science Dr., Newport, OR 97365-5260, (503) 867-4031 

In the bulk sediment toxicity test (BSTT) 

approach, test organisms are exposed in the 

laboratory to sediments collected in the field. To 

measure toxicity, a specific biological endpoint 

is used to assess the response of the organisms to 

the sediments. The bulk sediment toxicity approach is a descriptive method and cannot be 

used by itself to generate sediment quality criteria. 

3.1 SPECIFIC APPLICATIONS 

3.1.1 Current Use 

Sediment toxicity testing has been applied in 

dredged material disposal permit and other 

regulatory programs in the following ways 

(USEPA/USACE, 1991). 

n To determine potential biological hazards 

of dredged material intended for disposal 

in an aquatic environment; 

n To evaluate the effectiveness of various 

dredged material management actions; 

n To indicate the spatial distribution of 

toxicity in contaminated areas, the relative degree of toxicity, and the changes 

in toxicity along a gradient of pollution 

or with respect to distance from pollutant 

sources (Scott and Redmond, 1989; 

Swartz et al., 1982, 1985b); 

• To reveal temporal changes in toxicity 

(i.e., by sampling the same locations 

over time or by assaying layers of buried 

sediment in core samples) (Swartz et al., 

1986, 1991); 

• To reveal hot spots of contaminated sediment for further investigation (Chapman, 

1986); and 

• To rank sediments based on toxicity to 

benthic organisms and to define cleanup 

boundaries of small or large problem 

areas of contaminated sediment. 

BSTT integrates interactions among complex 

mixtures of contaminants that may be present in 

the field. Many classes of chemical contaminants, including metals, polycyclic aromatic 

hydrocarbons (PAHs), polychlorinated biphenyls 

(PCBs), dioxins, and chlorinated pesticides can 

contribute to toxicity in effluents and sediments 

(Chapman et al., 1982). The BSTT measures the 

total toxic effect of all contaminants, regardless 

of their physical and chemical composition. 

3.1.2 Potential Use 

By itself, BSTT cannot generate chemicalspecific toxic effects data, but it can determine 

toxicity. Used in conjunction with toxicity 

identification evaluation procedures (Ankley et 

al., 1990) such as those described in Chapters 5, 

10, and 11, BSTT could help identify causal 

toxicants. To generate sediment quality criteria, 

the procedure must be combined with other 

methods of estimating sediment quality such as 

the Triad (Chapman, 1986b; Chapman et al., 
Sediment Classification Methods Compendium 

1987; see Chapter 10) and the Apparent Effects 

Threshold (AET) approach (Tetra Tech, 1986; 

PTI, 1988; see Chapter 11). BSTT will be most 

valuable in verifying other methods used to 

develop sediment quality criteria. 

3.2 DESCRIPTION 

3.2.1 Description of Methods 

The toxicological approach involves exposing 

test organisms to sediments. The chemical composition of the sediments, which may be complex, 

need not be known. At the end of a specified 

time period, the response of the test organisms is 

examined in relation to a specified biological 

endpoint (e.g., mortality, growth, reproduction, 

cytotoxicity, alterations in development or respiration rate). Results are then statistically compared 

with control and reference sediment results to 

estimate sediment toxicity. 

3.2.1.1 Objectives and Assumptions 

The objective of BSTT is to derive toxicity 

data that can be used to predict whether the test 

sediment will be harmful to benthic biota. It is 

assumed that the behavior of chemicals in test 

sediments in the laboratory is similar to that in 

natural in situ sediments. The effects of various 

interactions (e.g., synergism, additivity, antagonism) among chemicals in the field or in dredged 

materials can be predicted from laboratory results 

without measuring total or bioavailable concentrations of potentially hundreds of contaminants in 

the test sediment (Swartz et al., 1989) and without 

a priori knowledge of specific pathways of interaction between sediments and test organisms 

(Kemp and Swartz, 1989). One of the strengths 

of this test is to integrate the effects of all contaminants. However, the effect of individual contaminants cannot be determined by BSTT, therefore 

limiting its use in source control. This method 

can be used for all classes of sediments and any 

chemical contaminants, but not to answer causeand-effect questions. 

3.2.1.2 Level of Effort 

Implementation of this procedure requires a 

moderate amount of laboratory effort. A variety 

of toxicity test procedures (see Methods below) 

have been developed and are fairly straightforward 

and well documented. 

3.2.1.2.1 Type of Sampling Required 

It is recommended that bulk sediments be 

collected for analysis of total solids, acid volatile 

sulfide, grain size, and total and dissolved organic 

carbon (ASTM, 1990a). Bulk and interstitial 

concentrations of chemicals of interest can be 

determined in subsamples of the sediment added 

to the toxicity test chambers to enhance the 

interpretation of toxicity results. However, methods for sampling interstitial water have not been 

standardized (ASTM, 199Ob). Sediment variables 

such as pH and Eh should also be monitored. 

3.2.1.2.2 Methods 

The American Society for Testing and Materials (ASTM) has developed standard guidelines for 

several BSTTs (ASTM, 1990a, 1991). The most 

commonly used of these partial life cycle tests 

feature the marine amphipods Rhepoxynius abronius, Eohaustorius estuarius, Ampelisca abdita, and 

Grandidierella japonica (ASTM, 1990a); the 

freshwater/estuarine amphipod Hyalella azteca 

(ASTM, 1990c); and the freshwater chironomid 

species Chironomus tentans and C. riparius 

(ASTM, 1990c). Brief generalized descriptions of 

these tests are given below. 

BSTTs with the two freshwater chironomid 

species are functionally very similar, differing 

only in the age of the organisms with which the 

test is initiated and the duration of the test. Both 

C. tentans and C. riparius are available from 

various aquatic toxicology laboratories and commercial sources, and both species are cultured 

easily in a laboratory setting. Toxicity tests are 

initiated by adding C. riparius <3 days old or C. 

tentans 10-14 days old (second instar) to test 

chambers that contain bulk sediment with overlying water in various ratios (e.g., 6 water:1 

3-2 
3-BS’ITApptoach 

sediment; Giesy et al., 1988). The length of the 

test also varies with the biologica endpoint of 

interest and the species used. If the biological 

endpoint of interest is growth and survival of the 

larvae, the test is terminated after lo-14 days by 

sieving the C. riparks or C. tentans from the 

sediment. It also is possible to conduct the test 

until the adults emerge, which will occur (depending on temperature) in approximately 30 days for 

C. riparks and 20-25 days for C. r~larts. Toxicity test procedures with C. riprius and C. 

tentans are given in more detail in Adams .et al. 

(1985), Nebeker et al. (1984), Giesy et al. (1988), 

Ingersoll and Nelson (1989), and ASTM (1991). 

Partial life-cycle toxicity tests with the freshwater/estuarine amphipod H. uzteca and bulk 

sediments have been conducted in a number of 

laboratories. H. azteca are available from various 

aquatic toxicology laboratories and commercial 

sources and can be cultured easily in a laboratory. 

Toxicity tests are initiated by adding juveniles ~7 

days old to test chambers that contain bulk sediment with overlying water in various ratios (e.g., 

4 water:1 sediment; Ingersoll and Nelson, 1989). 

‘Ihe length of the test can range from ~10 days 

(short-term partial life-cycle test) to 30 days (longterm partial life-cycle test) (Nebeker et a!,, 1984; 

Ingersoll and Nelson, 1989). Depending on the 

Iength of the test, biological endpoints include 

survival, behavior, growth, and reproduction. 

More detailed descriptions of toxicity test procedures are given by Nebeker et al. (1984), Nebeker 

and Miller (1988), Ingersoll and Nelson (1989), 

and ASTM (1991). 

Partial life-cycle toxicity tests with the marine 

amphipods Rhepoxynius abroniw, Eohaustorius 

estuarius, Ampelisca abdita, and Grandidierella 

japonica and bulk sediments have been used for 

some time (Swartz et al., 1985a). Amphipods and 

bulk sediments generally are collected from the 

field and acclimated to laboratory conditions for 

2-24 days before toxicity testing. The tests are 

initiated by adding immature or adult amphipods 

to test chambers that contain bulk sediment with 

overlying water in various ratios. The length of 

the test generally is ~10 days, and the biological 

responses monitored consist of behavioral effects 

(e.g., emergence from the sediment, ability to 

burrow in clean sediment after exposure to test 

sediment) and mortality. More detailed descrig 

t@ns of the toxicity test procedures are given by 

Swartz et al. (1985a), Dewitt et al. (1989), Nipper 

et al. (1989), Scott and Redmond (1989), ASTM 

(199Oa), and the Puget Sound Estuary Program 

(1991). Chronic test procedures for marine and 

estuarine amphipods are under development at 

several laboratories. Other test procedures for 

marine and estuarine polychaetes, pelecypods, 

shrimp, and fish are described in the USEPA/ 

USAGE (1991) and Reish and I&&y (1988) 

manuals for testing dredged materials before 

disposal. 

3.2.1.2.3 Types of Data Required 

The physical and chemical data described 

above under Section 3.2.1.2.1, Type of Sampling 

Required, are needed to interpret the test results. 

The required biological data (which vary by test) 

may include mortality and various sublethal 

effects (e.g., changes in growth, reproduction, 

respiration rate, behavior, or development). These 

data can be compared to control and reference 

data to determine the occurrence of biological 

effeds (ASTM, 199fla). Dilution experiments in 

which uncontaminated sediment is added to test 

sediment collected from the field can be used to 

calculate L&, values, EC, values, aocffcd 

concentrations, and lowest-observableeffect 

concentrations (Swark et al., 1989). 

3.2.1.2.4 Necessary Hardware and Skills 

In general, only re.adiIy available and inexpensive field ‘and laboratory equipment is needed, 

procedures are fairly simple and straightforward, 

and a minimum of training is necessary to detect 

endpoints through toxicity tests. Interpretation of 

the toxicity data (chemical and biological) requires 

a higher degree of skill and training. Chemical 

sampling methods are generally simple and rouhe, although analysis of chemical samples requires specialized training and equipment. Some 

biological effects tests also require specialized 

training, handling, and facilities. 

3-3 
Sediment Classifkdon Methods Compendium 

3.2.1.3 Adequacy of Documentation 

Various sediment toxicity test procedures have 

been developed and well documented for testing 

field sediments (A!XM 1990a; Chapman 1986a, 

1988; Lamberson and Swartz, 1988; Melzian, 

1990; Puget Sound Estuary Program (PSEP) 1991; 

Swartz, 1987; Thompson et al., 1989; USEPff 

USACE, 1991). Although standardization of 

methodology is progressing, intercalibration 

among laboratories and better field validation are 

needed. 

3.2.2 Applicability of Method to Human 

Health, Aquatic Life, or Wildlife 

Protection 

The BSlT approach is suitable only for 

protection of aquatic life. Sediment toxicity test 

procedures incorporate a direct measure of sediment biological effects and can be used to predict 

biological effects of contaminated sediments 

before approval of state and federal disposal 

permits. These procedures can be used to assess 

the Qxicity of sediments in the natural environment and to predict the effects of these sediments 

on resident aquatic life. Combined with other 

approaches such as the AET and the Triad 

approaches (Chapman, 1986b), BSTTs can be 

used to establish sediment quality criteria. Use of 

the most sensitive species within a benthic community as a test organism will serve to protect the 

structure and function of the entire ecosystem 

(Becker et al., 1990). 

3.23 Ability of Method to Genente 

Numerical Criteria for Specific 

Chemicals 

The BAIT approach cannot be used by itself 

to generate sediment quality criteria. Instead it 

must be combined with chemical measurements 

and other data to generate information on the 

effects of individual contaminants. Both the Triad 

and the AET approaches rely on bulk sediment 

toxicity data to derive numerical criteria. ms 

in conjunction with sediment quality criteria 

derived from equilibrium partitioning (USEPA, 

1980; Swartz et al., 1990) can also be used in 

assessments of potentially contaminated sediments 

(= Chapter 6, Equilibrium Partitioning 

Approach). 

33 USEFULNESS 

33.1 Environmental AppliabRkty 

3.3.1.1 Suitability for DiDrent Sediment l)pes 

The sediment toxicity test approach is suitable 

for any type of sediment. In some cases, the 

physical or chemical properties of the test sedimen4 such as salinity or grain size, may limit the 

selection of organisms that can be used for testing 

(Ott, 1986; Dewitt et al., 1989). Appropriate 

controls or statistical models (Dewitt et al., 1988) 

for sediment properties may be necessary to 

discriminate chemical toxicity from conventional 

effects. In establishing sediment quality criteria, 

the effects of features of the sediment itself, such 

as grain size, must be recognized (Dewitt et al., 

1988). Data can be normalized to such factors as 

organic carbon or acid volatile sulfide (DiToro et 

al., 1990, 1991; Nebeker et al., 1989) and thus 

can be applied to any sediment. However, normalization techniques are in the developmental 

stage (see Chapter 6, Equilibrium Partitioning 

Approach). 

3.3.1.2 Suitability for Different Chemicals or 

Classes of Chemicals 

BSTI’ is the only currently available approach 

that directly measures the biological effects of all 

classes of chemicals, including the combined 

.interactive (additive, synergistic, antagonistic) 

toxic effects among individual chemicals in 

mixtures of contaminants usually found in field 

sediments (Plesha et al., 1988; SW& et al., 

1989). Bioaccumulative chemicals can be evaluated if the length of the test is extended to ensure 

adequate exposure of the test organism. 

34 
3-“STT Appruach 

3.3.1.3 Suitability for Predicting Eficts on 

Diflerent Organisms 

Theoretically, any organism can be used in 

sediment toxicity testing. To protect a biological 

community and to predict the effects of contaminated sediments on different organisms, test 

organisms should be selected on the basis of their 

sensitivity to contaminants, their ability to withstand laboratory handling, and their ability to 

survive in control and reference treatments 

(Dewitt et al., 1989, Reish and LeMay, 1988; 

Shuba et al., 1981). In tests to determine the 

effects of contaminated sediments on a particular 

biological community, the test species selected 

should be among the most sensitive found in the 

community of interest, or should be comparably 

sensitive. Test species should include more than 

one type of organism to ensure a range of sensitivity to various types of contaminants (Becker et 

al., 1990). 

3.3.1.4 Suitability for In-Place Pollutant Control 

Sediment toxicity testing can be used directly 

to monitor in-place pollution. As discussed in 

Section 3.2.1.1, sediment toxicity testing can be 

used to determine the extent of the problem area, 

monitor temporal and spatial trends, detect the 

presence of unsuspected hot spots, assess the need 

for remedial actions, and monitor changes in 

toxicity after remediation. Such tests can also be 

used as a cost-effective and rapid screening tool 

for in situ pollutant reconnaissance surveys and in 

a priori simulations of proposed remedial actions 

to test the effectiveness of capping or other remedial alternatives. 

3.3.1.5 Suitability for Source Control 

Bulk field sediment toxicity testing can be 

used .to identify suspected sources of sediment 

pollution. Field reconnaissance surveys can reveal 

hot spots near contaminant sources, and a map 

showing contours of sediment toxicity values can 

reveal gradients that identify point and aonpoint 

sources (Swartz et al., 1982). Toxicity testing 

cannot be used by itself to verify reductions in the 

mass loading of chemicals that might be expected 

as a result of source control. However, the biological effects of source control can be represented 

through the use of BSTT. 

33.1.6 Sui&bility for Disposd Applicatims 

BSlT has been used widely in regulatory 

programs to determine the toxicity of material 

before disposal (Reish and LeMay, 1988; USEPA/ 

USACE, 1991). The potential hazard to benthic 

organisms at the disposal site (which is determined by making comparisons with the “reference” sediments collected near the disposal site) 

can be predicted from laboratory toxicity test 

results. Sediment toxicity tests also can be used 

to monitor conditions at the disposal site both 

before and after a disposal operation. 

3.33 General Advantages and Llmltatlons 

33.2.1 Ease of Use 

Most sediment toxicity test procedures are 

simple to use, requiring limited expertise and 

standard inexpensive laboratory equipment (PSEP, 

1991). Only a few sublethal effects tests require 

specialized training. Field sampling requires only 

readily available equipment and standard 

procedures (ASTM, 199Ob). 

3.3.2.2 Relative Cost 

Individual laboratory toxicity tests and field 

sampling are cost-effective because they require 

limited expertise and inexpensive equipment. 

Such costs generally range from $150 to $500 per 

sampling replicate. Laboratory sediment toxicity 

testing is a comparatively inexpensive and costeffective method of monitoring the field distribution of sediment toxicity because it integrates 

the effects of all toxic contaminants, does not 

require individual chemical measurements, and 

does not require time-consuming analysis of 

benthic community structure. 

3-5 
Sediment Cksificaficm Methods Compendium 

3.3.2.3 Tendency to Be Conservative 

Sediment toxicity tests can be made as sensitive or as conservative (i.e., environmentally 

protedive) as necessary through selection of 

biological endpoints and species of test organism. 

Reliance on mortality as an endpoint may be 

underprotective, while some sublethal endpoints 

(e.g., enzyme inhibition) may be overprotective. 

3.3.2.4 Level ofAcceptance 

B!STI’ is widely accepted by the scientific and 

regulatory communities and has been tested and 

contested in court. Field sediment toxicity test 

results have been published widely in peerreviewed journals and incorporated into other 

measures of sediment quality such as the AET 

and the Triad approaches. Standard guides for 

sediment toxicity testing continue to be developed 

by ASI’M (199Oa, 199Ob, 1991), and field sediment toxicity testing is incorporated into most 

dredged material disposal regulatory programs 

(PSEP, 1991; Reish and LeMay, 1988; USEPA/ 

USAGE, 1991). Toxicity testing in general has 

long been the basis for water quality criteria, 

dredged material testing, effluent testing, and 

discharge monitoring. 

3.3.2.5 Ability to Be Implemented by 

Laboratories with 7)pical Equipment 

and Handling Facilities 

Sediment toxicity test methods are easily 

implemented by laboratories with typical equipment using inexpensive glassware and procedures 

requiring little specialized training, although the 

interpretation of some sublethal biological endpoints may require some degree of training and 

experience. Field sediment sample collection 

procedures are routine. 

3.3.2.6 Level of Effort Required to Generate 

Results 

This procedure consists of field sampling and 

a laboratory toxicity test. Compared to an extensive survey of chemical concentrations or benthic 

community structure analysis, the level of effort is 

relatively small. 

3.3.2.7 Degree to Which Results Lend 

Themselves to Interpretation 

Biological responses to toxic sediment can be 

easily interpreted. Generally, data fit “pass-fail” 

criteria (i.e., the result is either above or below a 

predetermined acceptance level) or the result is 

compared statistically to control and reference 

results to determine whether there is a toxic effect. 

Xiltle expert guidance is required for interpretation 

of mortality data although chronic or sublethal 

effects might require some explanation. 

3.3.2.8 Degree of Environmental Applicability 

As noted in Section 3.3.1.1, the sediment 

toxicity test approach applies to a wide range of 

environmental conditions and sediment types. The 

effects of various sediment properties such as 

grain size and organic content can be addressed 

experimentally with appropriate uncontaminated 

controls. 

3.3.2.9 Degree of Accuracy and Prekion 

Because the sediment toxicity test is a laboratory-controlled experiment, its results have a high 

degree of accuracy, precision, and repeatability. 

3.4 STATUS 

3.4.1 Extent of Use 

Sediment toxicity tests are widely used in 

research and regulatory programs in both marine 

and freshwater systems (ASTM, 199Oa, 1991), as 

described in Section 3.2.1.1. Sediment toxicity 

tests also are incorporated into the evaluation of 

applications for dredged material disposal permits 

and are used to assess the toxicity of sediments 

subject to regulatory decisions. BSITs are used 

to investigate the mechanisms of sediment toxicity 

to benthic organisms (Kemp and Swartx, 1989; 

Swartz et ul., 1988). 

34 
3-BS’ITAppmch 

3.4.2 Extent to Which Approach I-hs Been 

Field-Validated 

Field validation of BSIT includes several 

publications in peer-reviewed literature (Chapman, 

1986b; Plesha et al., 1988; Swartz et al., 1982, 

1986, 1989). As more data become available, 

results can be compared with available information on contaminant concentrations in sediment in 

areas where biological effects have been observed. 

The effects of interactions among contaminants, as 

well as the effects of nonchemical sediment 

variables, must be taken into consideration when 

attempts are made at field validation (Dewitt et 

al., 1988; Swartz et al., 1989). h noted in 

Section 3.2.1.3, better field validation of predicted 

effects is needed. 

3.43 Reasons for Limlted Use 

BS’IT has been widely used in research and 

regulatory programs (see Section 3.4.1, Extent of 

Use). 

3.4.4 Outlook for Future Use and Amount of 

Development Yet Needed 

The outlook for future use of sediment toxicity tests is promising where direct measurement 

of biological effects of toxicants in sediments is 

desired especially where the effects of chemical 

interactions are of interest. Development and 

standardization of biological testing methods 

should continue, especiaIly for tests using species 

locally available in geographic areas that have not 

been represented such as tropical and arctic regions. More emphasis should be placed on the 

development of procedures to measure chronic 

effects. Methods should be compared and standardized among laboratories, and results should be 

field-validated to establish their ability to predict 

biological effects on populations and communities 

in the field. As more toxicity tests are conducted 

and the results subject to a quality assurance 

review, results should be compiled in a central 

database so that comparisons can be made among 

species, methods, and laboratories. 

3.5 REFERENCES 

Adams, W.J., R.A. Kimerle, and R.G. Mosher. 

1985. Aquatic safety assessment of chemicals 

sorbed to sediments. pp. 429-453. In: Aquatic 

Toxicology and Hazard Assessment: Proceedings of the Seventh Annual Symposium. 

ASTM STP 854. R.D. Cardwell, R. Purdy and 

R.C. Bahner (eds.). American Society for 

Testing and Materials, Philadelphia, PA. 

Ankley, G.T., A. Katko, and J.W. Arthur. 1990. 

Identification of ammonia as an important 

sediment-associated toxicant on the lower Fox 

River and Green Bay, Wisconsin. Environ. 

Toxicol. Chem. 9:313-322. 

ASTM. 199Oa. E 136790. Guide for conducting 

lo-day static sediment toxicity tests with 

marine and estuarine amphipods. In: Annual 

Book of ASTM Standards, Water and Enviroamental Technology, Vol. 11.04. American 

Society for Testing and Materials, Philadelphia, 

PA. 

ASTM. 1990b. E 139190. Guide for collection, 

storage, characterization and manipulation of 

sediments for toxicological testing. In: Annual 

Book of ASTM Standards, Water and Environmental Technology, Vol. 11.04. American 

Society for Testing and Materials, Philadelphia, 

PA. 

ASTM. 199Oc. E 138390. Guide for conducting 

sediment toxicity tests with freshwater invertebrates. In: Annual Book of ASTM Standards, 

Water and Environmental Technology, Vol. 

11.04. American Society for Testing and 

Materials, Philadelphia, PA. 

Becker, D.S., G.R. Bilyard, and T.C. Ginn. 1990. 

Comparisons between sediment bioassays and 

alterations of benthic macroinvertebrate assemblages at a marine Superfund site: Commencement Bay, Washington. Environ. Toxicol. 

C&em. 9: 669685. 

Chapman, P.M. 1986a. Sediment bioassay tests 

provide data necessary for assessm ent and 

regulation. In: Proceedings of the Eleventh 

Annual Aquatic Toxicology Workshop; Technical Report 1480. Green, G.H. and ILL 

Woodward (eds.). Fish. Aquat. Sci., pp. 178- 

197. 

3-7 
Sediment Cfa.ssijiuhon Methods Compendium 

Chapman, P.M. 1986b. Sediment quality criteria 

from the sediment quality triad: an example. 

Environ. Toxicoi. Chem. 5: 957-964. 

Chapman, P.M. 1988. Marine sediment toxicity 

tests. In: Chemical and Bioiogical Characterization of Sludges, Sediments, Dredge Spoils, and 

Drilling Muds, ASTM SIP 976, J-1. Iichtenberg, 

FA. Winter, C.I. Weber, and L Fradkin, (eds.). 

American Society for Testing and Materials, 

Philadelphia, PA pp. 391-402. 

Chapman, P.M., G.A. Vigers, MA. FarrelI, R.N. 

Dexter, E.A. Quinlan, R.M. &can, and M. 

Landolt. 1982. Survey of biological effects of 

toxicants upon Puget Sound biota. 1. Broadscale toxicity survey. NOAA Technical Memorandum OMPA-25, Boulder CO. 

Chapman, P.M., R.N. Dexter, and E. R. Long. 

1987. Synoptic measures of sediment contamination, toxicity and infaunal community composition (the sediment quality triad) in San 

Francisco Bay. Mar. Ecol. Prog. Ser. 37: 

75-96. 

Dewitt, T.H., G.R. Ditsworth, and R.C. Swartz. 

1988. Effects of natural sediment features on 

the phoxocephalid amphipod, Rhepoxynius 

obronius: Implications for sediment toxicity 

bioassays. Mar. Environ. Res. 25: 99-124. 

Dewitt, T.H., R.C. Swartz, and J.O. Lamberson. 

1989. Measuring the toxicity of estuarine sediments. Environ. Toxicol. Chem. 8: 1035-1048. 

DiToro, D.M., J.D. Mahony, D.J. Hansen, KJ. 

Scott, M.B. Hicks, S.M. Mayr, and M.S. 

Redmond. 1990. Toxicity of cadmium in sediments: the role of acid volatile sulfide. 

Environ. Toxicol. Chem. 9: 1487-1502. 

DiToro, D.M., J.D. Mahony, D.J. Hansen, KJ. 

Scott, A.R. Carlson, and G.T. Ankley. 1991. 

Acid volatile sulfide predicts the acute toxicity 

of cadmium and nickel in sediments. Environmental Science and Technology. 

Giesy, J.P., R.L. Graney, J.L. Newsted, C.J. 

Rosiu, A. Benda, R.G. Kreis, and F.J. Horvath. 

1988. Comparison of three sediment bioassay 

methods using Detroit River sediments. 

Environ. Toxicol. Chem. 7: 483-498. 

Ingersoll, C.G., and M.K. Nelson. 1989. Solidphase sediment toxicity testing with the freshwater invertebrates: Hyalella azteca (Amphipoda) and Chironomus riparius (Diptera). In: 

AquaticToxicology Risk Assessment: Proceedings of the Thirteenth Annual Symposium, 

ASTM m, American Society for Testing and 

Materials, Philadelphia, PA. 

Kemp, P.F., and R.C. Swartz. 1989. Acute toxicity 

of interstitial and particle-bound cadmium to a 

marine infaunal amphipod. Marine Environ. 

Res. 26: 135153. 

Lamberson, J.O., and R.C. Swartz. 1988. Use of 

bioassays in determining the toxicity of sediment to benthic organisms, Chapter 13, In: 

Toxic Contaminants and Ecosystem Health: A 

Great Lakes Focus. Evans, M.S. (ed.). John 

Wiley and Sons, New York, NY. pp. 257-279. 

Melzian, B-D. 1990. Toxicity assessment of 

dredged materials: acute and chronic toxicity 

as determined by bioassays and bioaccumulation tests. In: Proceedings of the International Seminar on the Environmental 

Aspects of Dredging Activities, Goubault 

Impremeur, Nantes; France, pp. 49-64. 

Nebeker, A.V., MA. Cairns, J.H. Gakstatter, KW. 

Maleug, G.S. Schuytema, and D.F. Kraw@c. 

1984. Biological methods for determining toxicity 

of contaminated freshwater sediments to invertebrates. Environ. Toxicol. C&m. 3: 617430. 

Nebeker, A-V. and C.E. Miller. 1988. Use of the 

amphipod crustacean Hyalefla artecu in freshwater and estuarine sediment toxicity tests. 

Environ. Toxicol. Chem. 7: 1027-1034. 

Nebeker, A.V., G.S. Schuytema, W.L Griffis, JA. 

Barb&a, and LA. Carey. 1989. Effect of 

sediment organic carbon on survival of 

Hyalellu uztecu exposed to DDT and endrin. 

Environ. Toxicol. and (3hem. 8: 705-718. 

Nipper, M.G., D J. Greenstein, and S.M. Bay. 

1989. Short- and long-term sediment toxicity 

test methods with the amphipocl Grandidtirellu 

jtiporrica. Environ. Toxicol. and Chem. 8:1191- 

1200. 

Ott, F.S. 1986. Amphipod sediment bioassays: 

Effect of grain size, cadmium, methodology, 

and variations in animal sensitivity on interpretation of experimental data. Ph.D. dissertation. 

University of Washington, Seattle, WA. 

Plesha, P.D., J.E. Stein, M.H. Schiewe, B.B. 

McCain, and U. Varanasi. 1988. Toxicity of 

38 
3-BS7-T Approach 

marine sediments supplemented with mixtures of selected chlorinated and aromatic 

hydrocarbons to the infaunal amphipod, 

Rhepoxynius abronius. Mar. Environ. Res. 

25: 85-97. 

Puget Sound Estuary Program. 1991. Recommended guidelines for conducting laboratory bioassays on Puget Sound sediments. Draft report 

prepared for U.S. Environmental Protection 

Agency, Region 10, Office of Puget Sound, 

SeatUe, WA. 

PTI Environmental Services. 1988. Sediment 

quality values refinement: Tasks 3 and 5 - 

1988 update and evaluation of the Puget Sound 

AET. PTI Environmental Services, Bellevue, 

WA. 

Reish, D.J., and J.A. Lemay. 1988. Bioassay 

manual for dredged materials. Contract 

DACW-O9-83R-005. U.S. Army Corps of 

Engineers, Los Angeles District, Los Angeles, 

CA. 

Scott, KJ., and M.S. Redmond. 1989. The effects 

of a contaminated dredged material on laboratory populations of the tubicolous amphipod, 

Ampelisca obdita. In: Aquatic Toxicology and 

Hazard Assessment: Vol 12. U. M. Cowgill 

and L. R. Williams (eds.). ASTM STP 1027. 

American Society for Testing and Materials, 

Philadelphia, PA. 

Shuba, PJ., S.R. Petrocelli, and R.E. Bentley. 

1981. Considerations in selecting bioassay 

organisms for determining the potential environmental impact of dredged material. Technical Report EL81-8. U.S. Army Engineer 

Waterways Experimental Station, Vicksburg, 

MS. 

Swartz, R. C. 1987. Toxicological methods for 

determining the effects of contaminated sediment on marine organisms. pp. 183-198. In: 

Fate and Effects of Sediment Bound Chemicals 

in Aquatic Systems. K. L. Dickson, A.W. 

Maki, and W. A. Brungs (eds.). Pergamon 

Press, New York. 

Swartz, R.C., WA. DeBen, K.A. Sercu, and J.O. 

Lamberson. 1982. Sediment toxicity and the 

distribution of amphipods in Commencement 

Bay, Washington, USA. Mar. Poll. Bull. 13: 

359-364. 

Swartz, R.C, WA. DeBen, J.K.P. Jones, J.O. 

Lamberson, and FA. Cole. 1985a. Phoxocephalid amphipod bioassay for marine sediment toxicity. In: Aquatic Toxicology and 

Hazard Assessment. R.D. Cardwell, R. Purdy 

and R.C. Bahner (eds.). ASIN STP 854, 

pp. 284-307. American Society for Testing 

and Materials, Philadelphia, PA. 

Swat-k, R.C., D.W. Schults, G.R. Ditsworth, WA. 

DeBen, and FA. Cole. 1985b. Sediment 

toxicity, contamination, and macrobenthic 

communities near a large sewage outfall. 

pp. 152-175. In: Validation and Predictability of Laboratory Methods for Assessing 

the Fate and Effects of Contaminants in Aquatic Ecosystems. T.P. Boyle (ed.). ASTM SIP 

865. American Society for Testing and Materials, Philadelphia, PA. 

Swartz. R.C., FA. Cole, D.W. Schults, and WA. 

DeBen. 1986. Ecological changes on the Pales 

Verdes Shelf near a large sewage outfall: 

1980-1983. Mar. Ecol. Prog. Ser. 31: l-13. 

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. Environ. 

Toxicol. Chem. 7: 1013-1020. 

Swartz, R.C., P.F. Kemp, D.W. Schults, G.R. 

Ditsworth, and R.O. Ozretich. 1989. Acute 

toxicity of sediment from Eagle Harbor, .Wasbington, to the infaunal amphipod Rhepoxymius 

abronius. Environ. Toxicol. and C%em. 8: 215 

222. 

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 equilibrium partitioning 

approach to sediment quality criteria. Environ. 

Toxicol. and Chem. 9: 1071-1080. 

Swartz, R.C., D.W. Schults, J.O. Lamberson, RJ. 

Ozretich, and J.K. Stull. 1991. A toxicological 

record in cores of contaminated sediment. 

Mar. Environ. Res. 

Tetra Tech, Inc. 1986. Eagle Harbor preliminary 

investigation. Final Report EGHB-2, TC3025003. Tetra Tech, Inc., Bellevue WA. 

Thompson, B.E.: S.M. Bay, J.W. Anderson, J.D. 

Laughlin, D.J. Greenstein, and D.T. Tsukada. 

3-9 
Sdiment Classijkdon Methds Compendium 

1989. (konic effects of contami-nated sediments on tbe urchin Lytichinuspictus. Environ. 

Toxicol. and Chem. 8: 629-637. 

USEPA. 1980. Water quality criteria for fluoranthene. U.S. Environmental Protection 

Agency, Washington, DC. 

USEPANSACE. 1991. Evaluation of dredged 

material proposed for ocean disposal-testing 

manual. EPA-5034%91/001. U.S. Environmental Protection Agency and U.S. Army 

Corps of Engineers, Washington, DC. 

3-10 
CHAPTER 4 

Spiked-Sediment Toxicity Test Approach 

Janet O. Lamberson end Richard C. Swartz 

U.S. Environmental Protection Agency, Pacific Ecosystems Branch, ERL-N 

2111 Southeast Marine Science Dr., Newport, OR 97365-5260 

(503) 867-4031 

The toxicological approach to generating sediment quality criteria uses concentration-response 

data from sediments spiked in the laboratory with 

known concentrations of contaminants. Sediments 

are spiked to establish cause-and effect relationships between chemicals and adverse biological 

responses (e.g., mortality, reduction in growth or 

reproduction, physiological changes). Individual 

chemicals or other potentially toxic substances can 

be tested alone or in combination to determine 

toxic concentrations of contaminants in sediment. 

This approach can be used to generate sediment 

quality criteria or to validate sediment quality 

criteria generated by other approaches. 

4.1 SPECIFIC APPLICATIONS 

4.1.1 Current Use 

The spiked-sediment toxicity test (SSTT) 

approach is in the research stage. Although the 

procedures used resemble those used to generate 

water quality criteria, the influence of the variable 

properties of sediment makes generating quality 

criteria values much more complex. 

Where LC50 values and chronic effects data 

are available for chemicals in sediments (see 

Section 4.3.2.3), they can be used to identify 

concentrations of chemicals in sediment that are 

protective of aquatic life. The predictive value of 

sediment quality criteria generated by this 

approach should be tested by comparing them 

with field data on chemical concentrations in 

natural sediments and observed biological effects. 

However, interim laboratory-derived criteria can 

be implemented before field validation. 

4.1.2 Potential use 

This method can be used to address empirically the problem of interactions among complex 

mixtures of contaminants that are almost always 

present in the field (Swartz et al., 1988, 1989). 

Chemical-specific data can be generated for a 

wide variety of classes of chemical contaminants, 

including metals, PAHs, PCBs, dioxins, and 

chlorinated pesticides. Both acute and chronic 

criteria can be established, and the approach is 

applicable to both marine and freshwater systems 

(Tetra Tech, 1986; Battelle, 1988). However, 

unless the sediment factor that normalizes for 

bioavailability is known, this procedure must be 

applied to every sediment (i.e., a value derived for 

one sediment may not be applied with predictable 

results to another sediment with different 

properties). 

4.2 DESCRIPTION 

4.2.1 Description of Method 

The toxicological approach involves exposing test organisms to sediments that have been 

spiked with known quantities of potentially toxic 

chemicals or mixtures of compounds. At the 

end of a specified time period, the response of 

the test organism is examined in relation to a 

biological endpoint (e.g., mortality, growth, 

reproduction, cytotoxicity, alterations in development or respiration rate). Results are then 

statistically compared with results from control 

or reference sediments to identify toxic concentrations of the test chemical. 
Sediment Classification Methods Compendium 

4.2.1.1 Objectives and Assumptions 

The objective of this approach is to derive in 

the laboratory concentration-response values that 

can be used to predict the concentrations of 

specific chemicals harmful to resident biota under 

field conditions. The effects of the interactions--synergism, additivity, antagonismamong chemicals in the field can be predicted 

from laboratory results with sediments spiked with 

combinations of chemicals. This method can be 

used for all classes of sediments and any chemical 

contaminant. The bioavailable component of 

contaminants in sediment can be determined by 

this method, and an a priori knowledge of specific pathways of interaction between sediments and 

test organisms is not necessary. Any method of 

expressing the bioavailability of contaminants in 

sediment can be used with sedimentoxicity tests, 

including the “free” interstitial concentration and 

normalization to organic carbon, acid volatile 

sulfide, and other sediment properties. 

Data generated by this method may be difficult to interpret if the normalizing factor for 

bioavailability is unknown. If the normalization 

factor is known, this method can be used to 

validate sediment quality criteria generated by 

other approaches. It is assumed that laboratory 

results for a given sediment and overlying water 

represent biological effects of similar sediments in 

the field, and that the behavior of chemicals in 

spiked sediments is similar to that in natural, in 

situ sediments. 

4.2.1.2 Level of Effort 

Implementation of this procedure requires a 

moderate to considerable amount of laboratory 

effort. The various toxicity test procedures that 

have been developed are generally straightforward and well documented (Lamberson and 

Swartz, 1988; Melzian, 1990; Nebeker et al., 

1984; Swartz et al., 1989; PSEP, 1991). However, many individual tests would be required to 

generate an extensive database of sediment quality 

values for a large number of chemicals, chemical 

combinations, and sediment types. 

4.2.1.2.1 Type of Sampling Required 

Collection of sediments from the field is 

required. Depending on the particular study 

objectives, the sediments may be clean (uncontaminated) sediments from a control area, uncontaminated reference sediments for comparison with 

similarly contaminated sediments, or contaminated 

sediments to be spiked with known concentrations 

of chemicals in a test for interactions among 

contaminants. Sufficient sediment must be collected to provide samples for chemical analysis, 

spiking, and reference or controls (i.e., sediment 

for statistical comparison with spiked sediment). 

Depending on the experimental design, the following controls might be required: sediment from the 

collection site for test animals (or culture sediment 

for laboratory-cultured animals), positive controls 

with a reference toxicant, carrier controls, and 

reference sediment controls for natural sediment 

features that may affect test animals, such as grain 

size distribution (Dewitt et al., 1988). 

4.2.1.2.2 Methods 

Various methods of adding chemicals to 

sediment (spiking sediments) have been used. In 

general, the chemical is either added to the sediment and mixed in (Birge et al., 1987; Ditsworth 

et al., 1990; Francis et al., 1984) or added to the 

overlying water (Hansen and Tagatz, 1980; Kemp 

and Swartz, 1988) or to a sediment slurry (Landrum, 1989; Oliver, 1984; Schuytema et al., 1984) 

and allowed to equilibrate with the sediment. 

Sediments are spiked with a range of concentrations to generate LC50 data or to determine a 

minimum concentration at which biological effects 

are observed. 

The effect of sediment contaminants on 

benthic biota is determined either by exposing 

known numbers of individual benthic test organisms to the sediment for a specific length of time 

(Swartz et al., 1985) or by exposing larvae of 

benthic species to the sediment in flowing natural 

waters (Hansen and Tagatz, 1980). Biological 

responses are determined at the end of the test 

period using response criteria that include mortality, changes in growth or reproduction, behavioral 

4-2 
4-SSl-T Apptoach 

or physiological alterations, or differences in 

numbers and species of larvae in contaminated 

versus control sediments. 

4.2.1.2.3 Types of Data Required 

Spiked sediments, as well as reference or 

control sediments, must be analyzed for total 

solids, grain size, and total and dissolved organic carbon. The concentralions of toxicants 

added to sediment must be determined ‘in stock 

solutions as well as in the test sediment. Bulk 

and interstitial levels of the spiked chemicals in 

the test sediment must be determined throughout 

a concentration range at least at the beginning 

and at tbe end of the toxicity test. However, 

methods for sampling interstitial water have not 

been standardized. If sediment properties lbat 

control availability, such as acid volatile sulfides 

or dissolved or total organic carbon, change 

during exposure, measurements must be taken 

before, during and at the end of the exposure 

period. In addition, these changes must be taken 

into account in interpreting the data. Sediment 

parameters such as pH and Eh should also be 

monitored. 

Biological and chemical data are compared 

statistically with control or reference data to 

determine the occurrence of biological effects, 

and can be used to calculate LC, values, EC, 

values, no-effect concentrations, or lowestobservable-effect concentrations. Establishment 

of the maximum acceptable toxicanl concentration requires data from a chronic or life-cycle 

test. 

Data correlating observed biological effects 

with chemical concentrations in spiked sediment 

can be used to calculate probit curves for derivation of biological effect level values (e.g., EC,). 

Data from several species of lest organisms can 

be ranked, and the lowest contaminant concentrations that affect the most sensitive species can 

be used to establish sediment quality criteria that 

will protect tbe entire benthic community and 

associated aquatic ecosystem. This approach has 

regulatory and scientific precedence in the 

development of water quality criteria. 

4.2.1.2.4 Necessary Hardware and Skills 

Most toxicity test procedures require a minimum of specialized hardware and level of skill. 

In general, only readily available and inexpensive 

laboratory equipment is needed, procedures are 

fairly simple and straightforward, and a minimum 

of training is necessary to detect and interpret 

biological endpoints. Although analysis of chemical samples requires specialized training and 

equipment, the chemical sampling methods for 

spiked-sediment toxicity are generally simple and 

routine. Some biological effects tests also require 

specialized training and experience, especially to 

interpret the results. 

4.2.1.3 Adequacy of Documentation 

Various acute sediment toxicity test procedures have been developed and are well documented for testing freshwater and marine field 

sediments (Chapman, 1986,1988; Lamberson and 

Swartz, 1988; Melzian, 198% Swartz, 1987). 

Although only a few of these procedures have 

been used witb laboratory-spiked sediments, most 

of the established methods could be used with 

laboratory-prepared sediments as well as with field 

sediments. 

In contrast to acute tests, there are relatively 

few procedures for testing the chronic effects of 

contaminated sediments on benthic invertebrates. 

Life-cycle test methodology has been presented 

for the amphipods Ampelisca abdita (Scott and 

Redmond, 1989), HyaleZIa azteca (ASTM, 199oc, 

Borgmann and Munawar, 1989), and Grandidierella lutosa and G. lignorum (Connell and 

Airey, 1982); the polychaetes Neanthes arenaceo- 

&ntata (Pesch, 1979) and Capitellu capitata 

(Chapman .and Fink, 1984); freshwater oligochaetes (Wiederholm el al., 1987); and species of 

Daphnia and Chironomus (ASTM, 1991; Nebeker 

et al., 1988). Chronic exposures to most sensitive 

life stages are also inherent in the benthic recolonization procedure (Hansen and Tagatz, 1980). 

Further research is needed to develop and validate 

methodology for other species. 

4-3 
Sediment Class+cat~ Methods Compendium 

4.2.2 AppUabUty ot Method to Humw~ 

Health, Aquatic Ufe, or Wildlife 

Protectkul 

Spiked-sediment toxicity tests incorporate a 

direct measure of sediment biological effects. This 

approach is the only method that can quantify the 

interactive effects of combinations of contaminants 

diEtly. 

When chemical concentrations in tested biota 

are measured after a spikedsediment toxicity test, 

uptake of contaminants by benthic organisms (bioaccumulation) can be predicted. As an important 

component of food webs in aquatic ecosystems, 

benthic organisms can contriiute toxicants accumulated from contaminated sediments to higher levels 

of the aquatic food web and ultimately affect human 

health. Sediment quality criteria and bioauxrmulation studies using sediment toxicity test methods can 

help to set limits on the disposal of toxic sediments 

and predict uptake of toxicants into food webs. If 

this approach is combined with chemical analysis of 

sediment samples and BSTT, these limits can be 

used to defme areas from which food species should 

not be harvested or consumed or where direct 

contact with contaminated sediments can be hazardous to human health. 

Bioaccumulation studies and sediment quality 

uiteria estabkhed using data from SSTT with 

several benthic species can also be used to protect 

benthic communities and aquatic species that feed 

on the benthos. Assuming that a sufficient mix of 

taxonomic groups is used, a sediment quality aiterion based on the responses of the most sensitive 

species within a benthic community can be developed. This criterion can then be employed to 

protect the stnrcture and function of the entire 

ecosystem (Hansen and Tagatz, 1980). 

4.23 Ability of Method to Generate NumericA 

Criteria for Specific Chemkals 

Laboratory tests with the SSlT approach can be 

used to measure the effects of specific chemicals in 

various types of sediments directly and to establish 

unequivocal analysis of causal effeds. Test conditions allow this method to determine the effects of 

individual chemicals or mixtures of chemicals on 

44 

benthic biota (Plesha et al., 1988; SwarQ cl d., 

1988, 1989), establish pathways of toxicity, and 

provide specific effects concentrations (e.g. & 

EC, noeffect concentration). The influence of 

various physical &.rac&ristics of the sediment on 

cimnkal toxicity also can be determined (Dewitt er 

al., 1988, Ott 1986). ‘Ibe available data reprezznt 

concentrations at which toxicity occurs rather than 

numerical sediment quality criteria. Recent spiked 

sediment studies have provided data that can be 

useful in setting preliminary sediment criteria levels 

based on quiliirium partitioning models and water 

quality values (Swartz et al., 1990). 

Concentration-response data have been generated using SSIT for a variety of chemicals, including metals and organic compounds. Specific data 

are available for phenanthrene, fluoranthene, tic., 

mercury, copper, cadmium, hexachlorobenzene, 

pentachlorophenol, Aroclor 1242 and X54, chlordane, DDE, DDT, die&in, endosulfan, e&in, 

sevin, ueusote, and kepone (Adams et al., 1985; 

Cairns et al., 1984, Dewitt et al., 1989; Kemp and 

Swarlz, 1989; McLeese and MetcaRe, 1980; McIxese ei al., 1982; Nebeker et al., 198% Swartz et 

al., 1986, 1988, 1989; Tagatz et al., 1977, 1979, 

1983, Word et ol., 1987). Concentrations of nonionic organic compounds are usually normalized to 

sediment organic carbon or acid volatile sullide 

(D~Toro et al., 1990, 1991; Nebeker et al., 1989). 

Normalizing factors for other compounds in 

sediment currently are being researched. 

43 usEFl?LNEss 

43.1 Eovbwmental Appltcrbility 

4.3.1.1 Suitabihiy fbr Difirent Seahent Types 

The SSlT approach is suitable for any type of 

sediment. This appioach also can be used to 

establish the bioavailable component of the sediment responsible for the observed toxicity. The 

effects of various physical properties of the sediment on chemical toxicity can be determined 

experimentally. In some cases, the physical or 

chemical properties of the test sediment such as 

salinity or grain size may limit the species of 

organisms that can be used for testing, and a 
substitute species must be used (Dewitt et al., 

1988, 1989). When establishing sediment quality 

criteria, the effects of adverse physical or chemical 

properties of the sediment itself must be reflected. 

When factors controlling bioavailability (e.g., 

organic carbon, acid volatile sulfide) are known, 

data can be normalized to such factors, and the 

approach applied to any sediment type. 

4.3.1.2 Suitability for Different Chemicals or 

Classes of Chemicals 

A major advantage of the SSTI’ method is 

that it is suitable for all classes of chemicals. In 

addition, it is the only approach currently available that can empirically determine the interactive 

effects among individual chemicals in mixtures of 

contaminants usually found in real-world sediments (Swartz ef al., 1988, 1989). This approach 

also can be used to provide experimental validation of sediment quality criteria generated by other 

approaches. 

4.3.1.3 Suitability for Predicting E,@xts on 

Different Organisms 

Theoretically, any organism can be used in 

SSTT. To protect a biological community and to 

predict the effects of a toxicant on different organisms, test organisms should be selected based on the 

following criteria: (1) their sensitivity to contaminants, (2) their ability to withstand laboratory 

handling, and (3) their ability to survive in control 

treatments. Tests to determine the effeds of 

toxicants on a particular biological community 

should use the most sensitive species found in the 

community or a species with comparable sensitivity. 

4.3.1.4 Suhbifity for In-Place Pollutant Control 

SSIT can be used to develop sediment quality 

criteria, which will then be used to determine the 

extent of the problem area. It also can be used to 

monitor temporal and spatial trends and to assess the 

need for remedial action. Criteria can be used in 

setting target cleanup levels and in post-cleanup 

monitoring of actual contaminant levels. 

4.3.2.5 S~iliiy for Source Ctmbvl 

!SlT can be combined with wasteload aUocation models and used in source control to establish 

maximum allowable effluent concentrations or mass 

loadings of single chemicals and mixtums of drank 

cals. 

4.3.1.6 Suitabildy for Dis- Apphztians 

SSTT can be used to predid the biological 

effects of aMaminants before approval of dredged 

material disposal or sewage outfall permits. 

43.2 Genend Advantages md Unltatk~ns 

4.3.2.1 Ease of Use 

Most sediment toxicity test procedures arc 

simple to use, require limited expertise, and use 

standard laboratory equipment. Some of the sublethal-effects tests require specialized training. 

4.3.2.2 Relative Cost 

The cost of individual toxicity tests is relativeIy 

low because such tests require limited expertise and 

inexpensive equipment. (See Chapter 3, Bulk Se& 

ment Toxicity Approach.) ‘Ibe costs to implemtnt 

this approach as a regulatory tool would be ccmquatively high because SSIT requires the cokction of 

sediment chemistry data for comparison to data 

establiied by the sediient toxicity test method 

‘he cost of developing a large toxic&gical database would be relatively high because of the large 

number of individual chemicaIs and sediments that 

would have to be tested. Generating the citemicaI 

and toxicologicaI data necessary to establish I 

sediment quality criterion for one chemical by this 

method is estimated to cost $100,000. 

4.3.23 Tendency b Be Conservutiw 

L&oratorycontrolled SSIT experiments pmvide a high degree of accuracy. The tests are 

controlled sufficiently to give an estimate of the 

toxicity of individual chemicals in sediment. I.& 

oratory bioassays, especially acute toxicity tests, are 

4-s 
Se&nent Classification Methods Compmdium 

inherently limited in their ability to reflect ail of the 

ecological processes through which sediment contaminants may affect benthic ecosystems in the field. 

4.3.2.4 Level of Acceptance 

SSTT methods, which follow the procedures 

and rationale used to develop water quality criteria, are easily interpreted, technically acceptable, 

and legally defensible. The procedures and 

resulting data have been accepted and published in 

peer-reviewed journal articles, and some procedures have been incorporated into standard guidelines by ASTM’s subcommittee on sediment 

toxicology (ASTM, 19!3Oa, 199Oc). 

4.3.2.5 Abiliry to Be implemented by 

Laboratories with Qpical Equipment 

and Handling Facilities 

SSIT methods are implemented easily by 

laboratories with typical equipment, requiring 

inexpensive glassware and little specialized training. Spiking sediments may require special 

handling facilities for preparing stock solutions of 

hikhiy toxic substances, and the interpretation of 

some sublethal biological endpoints may require 

some degree of training and experience. 

4.3.2.6 Level of Eflort Required to Generate 

Results 

This procedure consists of a laboratory toxicity test and requires a moderate amount of effort 

to begin and end an experiment. The data generated must be compiled, and some calculations 

must be made to derive concentration-response 

relationships. The generation of chemical and 

biological data required for a large database of 

sediment quality values based on this approach 

would require a relatively high level of effort. 

4.3.2.7 Degree to Which Results Lend 

Themselves lo Interpretation 

Sediment toxicity tests applied to spiked sediments provide an unequivocal analysis of causeand-effect relationships between toxic chemicals 

4-4 

and biological responses. Because the procedures 

follow the rationale used in the development of 

water quality criteria, the methods are legally 

defensible. Toxicity tests bave long been accepted 

by both the public and the scientific community as 

a basis for water quality criteria and dredged 

material testing. 

4.3.2.8 Degree of Environmen la1 Applicability 

The SSIT approach is applicable to a wide 

range of environmen.tai conditions and sediment 

types. The confounding effects of sediment variables such as grain size and organic content can 

be addressed experimentally by using toxicity test 

methods or can be addressed by using normaiization equations (Dewitt et al. 1988). A major 

advantage of SSTI’ is the ability to predict interactive effects of chemical mixtures such as those 

found in field sediments. 

4.3.2.9 Degree of Accuracy and Precision 

Because the SSTT is a laboratory-controlled 

experiment, results have a high degree of accuracy 

and precision. The procedure produces a direct 

dose-response data set for individual chemicals in 

sediment. Sediment criteria generated by this 

approach must be field-validated. 

4.4 STATUS 

4.4.1 Extent of Use 

SSIT procedures are under development in 

several laboratories. Spiking procedures, as well 

as biological test procedures, are currently being 

standardized by ASTM’s sediment toxicology 

subuxnmittee (ASTM, E9Ob). 

4.4.2 Extent to Wbkb Approach Has Been 

Field-Validated 

Although some results have been published, 

spiked-sediment toxicity test values have not been 

well validated in the field, (Piesha et al., 1988; 

Swartz et al., 1989). As more data and criteria 
4-SSTT Approach 

values become available, they can be compared 

with existing information on contaminant levels in 

sediment in areas where biological effects have 

been observed. The effects of interactions among 

contaminants, as well as the effects of nonchemical sediment variables, must be considered during 

field validation (Dewitt et al., 1988; Swartz et al., 

1989). 

4.43 Reasons for Limited Use 

Although some data have been generated and 

compared to field conditions, the approach is still 

in the developmental stage in several laboratories, 

and a relatively large expenditure of effort will be 

needed to generate a large database. To date, 

there have been few comparisons of methods and 

species sensitivity, and few chronic toxicity tests 

have been developed. 

4.4.4 Outlook for Future Use and Amount of 

Development Yet Needed 

The outlook for future use of SSlTs or other 

sediment toxicity tests is promising where 

accurate, direct dose-response data are desired, or 

where the effects of chemical interactions need to 

be examined. Development of sediment-spiking 

and biological-testing methods should continue, 

methods should be compared and standardized 

among laboratories, and results should be fieidvalidated to establish their ability to predict 

biological effects in sediments. As more toxicity 

tests are conducted, results should be compiled 

in a central database so that comparisons can 

be made among species, methods, and laboratories 

and so that sediment quality criteria can be 

developed. 

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Seventh Annual Symposium. R.D. Cardwell, 

R. Purdy and R.C. Bahner (eds.). ASI’M SIT 

854, American Society for Testing and Materials, Philadephia, PA. 

Swartz, R.C, G.R. Ditsworth, D.W. Schults, and 

J.O. Lamberson.. 1986. Sediment toxicity to a 

marine infauna amphipod: cadmium and its 

interaction with sewage sludge. Mar. Environ. 

Res. 18: 133-153. 

Swartz, R.C., P.F. Kemp, D.W. schults, and J.O. 

Iamberson. 1988. Effects of mixtures of 

sediment contaminants on the marine infaunal 

amphipod, Rkepoqvks abro&s. Environ. 

Toxicol. Chem. 7: 1013-iO20. 

Swartz, R.C., P.F. Kemp, D.W. Schults, G.R. 

Ditsworth, and R.J. Ozretich. 1989. Toxicity 

of sediment from Eagle Harbor, Washington to 

the infaunal amphipod, Rhepoxyniw abronius. 

Environ. Toxicol. C&em. 8: 215-222. 

!%vartz, 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 equilibrium partitioning 

approach to sediment qoality criteria. Environ. 

Toxicol. Chem. 9: 1071-1080. 

Tagatz, M.E., J.M. Ivey, and H.K. Lehman. 1979. 

Effects of sevin on development of experimental estuarine communities. J. Toxicol. and 

Environ. Health 5: 643651. 

Tagatz, M.E., J.M. Ivey, J.C. Moore, and M. 

Tobia. 1977. Effects of pentachlorophenol on 

the development of estuarine communities. J. 

Toxicol. and Environ. Health 3: 501-506. 

Tagatz, M.E., G.R. Plaia, C.H. Deans, and E.M. 

Lores. 1983. Toxicity of aeosote-contaminated sediment to field- and laboratory-colonized 

estuarine benthic communities. Environ. 

Toxicol. cbem. 2: 441450. 

Tetra Tech, Inc. 1986. Development of sediment 

quality values for Puget Sound. Task 6 Final 

Report. Tetra Tech, Inc., Bellevue, WA. 

Wiederholm, T., A.M. Wiederholm, and G. 

4-9 
Sediment Chssificatbn Methods Compendium 

Milbrink. 1987. Bulk sediment bioassays with the toxicity of the nonpolar organic compound 

five species of fresh-water oligochaetes. DDT to the sediment dwelling organism 

Water, Air ahd Soil Pollut. 36: 131-154. Rhepoxynius abroth. Report prepared for U.S. 

Word, J.Q., JA. Ward, LM. Franklin, V.I. Environmental Protection Agency, fit&a and 

Cullinan and S.L. Kiesser. 1987. Evaluation of Standards Division, Battelle Washington Environthe equilibrium partition theory for estimating mental Program Office, Washington, DC. 
CHAPTER 5 

Interstitial Water Toxicity Identification 

Evaluation Approach 

Gerald Ankley and Nelson Thomas 

U.S. Environmental Protection Agency, Environmental Research Laboratory 

6201 Congdon Boulevard, Duluth, MN 55804 

(218) 720-5702 

The interstitial water toxicity approach is a 

multiphase procedure for assessing sediment toxicity 

using interstitial (pore) water. The use of pore 

water for sedimentoxicity assessment is based on 

the strong correlations between contaminant concentrations in pore water and observed exposure of 

benthic macroinvertebrates tosediment-associated 

contaminants (Adams et al., 1985; Swartz el al., 

1985; 1988; 1990, Connell et al., 1988; Knezovich 

and Harrison, 1988; USEPA, 1989a; DiToro et al., 

1990), as well as correlations between the actual 

toxicity of pore water and bulk sediments to epibenthic or benthic species (Ankley et al., 1991a). 

The approach combines the quantification of pore 

water toxicity with toxicity identification evaluation 

(TIE) procedures to identify and quantify chemical 

components responsible for sedimentoxicity (U.S. 

Environmental Protection Agency, 1988; 1989b; 

1989c, 1991a). TIE involves the use of toxicitybased fractionation procedures to identify toxic 

compounds in aqueousamples containing mixtures 

of chemicals (Burkhard and Ankley, 1989). In the 

interstitial water toxicity method, TIE procedures are 

implemented in three phases to characterize the 

nature of the pore water toxicant(s), identify the 

suspect toxicant(s), and confirm identification of the 

suspect toxicant(s). 

5.1 SPECIFIC APPLICATIONS 

5.1.1 Current Use 

The TIE procedures described herein were 

developed over the last 4 years using municipal and 

industrial effluents from more than 50 locations, as 

well as sediment samples from more than 10 different sites. They have been used with several aquatic 

species including cladocerans, fishes, and epibenthic 

macroinvertebrates. Although the methods were 

developed largely with freshwater species, they are 

generally applicable to, and are currently being used 

with, marine organisms as well. The procedures 

have proven to be successful in identifying acutely 

toxic substances in more than 90 percent of the 

samples to which they have been applied (e.g., 

Ankley et al., 1990a, 1991b; Kuehl et al., 1990; 

Amato et al, 1991; Norberg-King et al., 1991; 

Schubauer-Berigan and Ankley, 1991; Ankley and 

Burkhard., 1992). 

5.1.2 Potential uses 

The use of pore water as a fraction to assess 

sediment toxicity, in conjunction with associated 

TIE procedures, can provide data concerning specific compounds responsible for toxicity of contaminated sediments. These data could be critical to the 

success of remediation of toxic sediments, including 

the control of inputs of contaminants. 

In spite of existing uncertainties in preparing 

and using pore water to assess sediment toxicity, the 

ability to identify specific toxicants responsible for 

acute toxicity in contaminated sediments makes pore 

water an important test fraction. Thus this method, 

in conjunction with other sediment classification 

methods, could prove to be extremely valuable. 

5.2 DESCRIPTION 

5.2.1 Description of Method 

The interstitial water toxicity method involves 

three major steps: 

l Isolation of pore water from sediment 

samples; 
Sediment Classification Methods Compendium 

• Performance of toxicity tests on pore 

waters; and 

• Application of TIE procedures to pore 

water fractions. 

Pore water can be isolated from sediment 

samples by compression (squeezing) techniques, 

displacement of water from sediment via the use 

of inert gases, centrifugation, extraction via dialysis, and micro-syringe sampling (Knezovich et al., 

1987; Knezovich and Harrison, 1988; Sly, 1988; 

USEPA, 1991b). The most representative pore 

water samples may be obtained using the latter 

two procedures. However, the resulting sample 

volumes are too small to be useful for toxicity 

tests and associated TIE work. Centrifugation has 

been used in a number of studies evaluating the 

toxicity of sediment pore water (Giesy et al., 

1988; Swartz et al., 1989; Hoke et al., 1990; 

Ankley et al., 1990a; Schubauer-Berigan and 

Ankley, 1991) and comparative studies at Duluth, 

as well as other laboratories, indicate that centrifugation is a reasonable technique for pore water 

preparation (Schults et al., 1991; U.S. Environmental Protection Agency, 1991b). Regardless of 

the techniques chosen for pore water isolation, the 

method should not involve filtration either during 

or after preparation (Schubauer-Berigan and 

Ankley, 1991; USEPA, 1991b). 

After preparation of pore water, toxicity tests 

can be performed using either standard test species 

(e.g., USEPA, 1985a, 1985b) or various types of 

epibenthic or benthic organisms amenable to 

toxicity testing in aqueous samples (Ankley et al., 

1991a; USEPA, 1991b). In samples exhibiting 

acute toxicity, it is then possible to directly apply 

the TIE procedures described below in Section 

5.2.1.2.2. 

5.2.1. IObjectives and Assumptions 

The objective of the interstitial water toxicity 

method is to derive chemical-specific toxicity data 

in the laboratory that can be used to assess sediment toxicity in field situations. With this approach, it is possible to quantify toxicity in a 

sample and potentially to identify chemical comportents responsible for toxicity. The major 

assumption in using this method is that the ampounds that are toxic to test organisms in the pore 

water, as it is isolated in the laboratory, are the 

same compounds that cause toxicity in sediments 

in situ. 

5.2.1.2 Level of Effort 

Implementation of this method requires a 

moderate amount of laboratory effort, both to 

perform toxicity tests and to conduct TIE studies. 

The effort expended in the TIE studies will be 

proportional to the complexity of analyses required for the identification of suspected toxicants. 

5.2.1.2.1 Type of Sampling Required 

Bulk sediment must be obtained and pore 

water prepared from the sediments. Routine 

measurement of certain chemical components of 

the pore water should be conducted. These 

measurements should include (but are not limited 

to) pH, hardness, alkalinity, salinity (where appropriate), dissolved oxygen, sulfides, and ammonia. 

Certain of these variables, in particular pH, also 

should be monitored in the bulk sediment. 

5.2.1.2.2 Methods 

The framework for existing TIE procedures is 

summarized below. Greater detail (e.g., with 

respecto all possible results that could be generated) is available elsewhere (USEPA, 1988, 

1989b, 1989c), as are specific methods for performing sediment TIEs (USEPA, 1991b). 

Toxic sediment samples can potentially contain thousands of chemicals, and usually only a 

handful are responsible for the observed toxicity. 

The goal of the TIE method is to identify quickly 

and cheaply the chemicals causing toxicity. 

However, components causing toxicity can vary 

widely, and potential toxicants include cationic 

metals, polar and nonpolar organics, and anionic 

inorganics, as well as ammonia or hydrogen 

sulfide. In addition, when multiple toxicants are 

present, it must be possible to determine the 

proportion of the overall toxicity due to each 

toxicant. 

5-2 
S-Interstitial Water TIE Approach 

After preparation of pore water and performance of initial toxicity tests, the first step in the 

TIE is to separate toxic from nontoxic components 

in the pore water sample. To isolate the toxicants, 

sample manipulations and subsequent fractionation 

techniques are used in combination with toxicity 

tests (toxicity tracking). Each fractionation step 

consists of manipulations to identify the physical/cbemical properties of the sample toxicants, 

thereby enabling selection of the “correct” analytical technique for detecting, identifying, and 

quantifying the toxicants in the manipulated 

samples. Because there may be significantly 

fewer chemical components in the manipulated 

samples than in the original sample, the task of 

deciding which component is causing the toxicity 

is much easier. The toxicity-based TIE approach 

enables direct relationships to be established 

between toxicants and measured analytical data 

because toxicants are tracked through all sample 

fractionations, using the most relevant detector 

available, the organism. Establishing this relationship ultimately results in highly efficient TIES. 

With the toxicity-based TIE approach, detection of synergistic and antagonistic interactions, as 

well as matrix effects, for the toxicants is possible 

via toxicity tracking. A priori knowledge of the 

toxicants’ behavior in the aqueous phase is not 

required. 

The TIE approach is divided into three phases. 

Phase I consists of methods to identify the physical/chemical nature of the constituents causing 

acute toxicity. Phase II describes fractionation 

schemes and analytical methods to identify the 

toxicants, and Phase III presents procedures to 

confirm that the suspected toxicants are the cause 

of toxicity. 

Phase I: Toxicant Characterixatlon-In Phase 

I, the physical/chemical properties of toxicants are 

characterized by performing manipulations to alter 

or render biologically unavailable generic classes 

of compounds with similar properties. Toxicity 

tests, performed in conjunction with the manipulations, provide information on the nature of the 

toxicants. Successful completion of Phase I 

occurs when both the nature of the components 

causing toxicity, as well as their consistent presence in a number of samples, can be established. 

After Phase I, the toxicants can be tentatively 

categorized as having chemical characteristics of 

cationic metals, nonpolar organics, polar organics, 

volatiles, oxidants, and/or substances whose 

toxicity is pHdependent. 

An overview of the sample manipulations 

employed in Phase I is shown in Figure S-l. Not 

shown in Figure 5-1, but performed on all samples, are routine water chemistry measuremats 

including pH, hardness, conductivity, and dissolved oxygen. These routine measurements are 

needed for designirlg sample manipulations and 

interpreting test data. The manipulations shown in 

Figure 5-1 are usually sufficient to characterize 

toxicity caused by a single chemical. When 

multiple toxicants are present, various combinations of the Phase I manipulations will most likely 

be required for toxicant characterization. 

Many of the manipulations in Phase I require 

samples that have been pH-adjusted. The adjustment of pH is a powerful too1 for detecting cationic and anionic toxicants since their behavior is 

strongly influenced by PH. By changing pH, the 

ratio of ionized to un-ionized species in solution 

for a chemical is changed significantly. The 

ionized and un-ionized species have different 

physical/chemical properties as well as toxicities. 

In Phase I, pH manipulations are used to examine 

two different questions: 

l Is the toxicity different at various pHs? 

. Does changing the pH, performing a 

sample manipulation, and then readjusting 

to ambient pH affect toxicity? 

The graduated pH test examines the first question, 

and the pH adjustment. aeration, filtration, and 

solid phase extraction (SPE) manipulations examine the second. 

In the graduated pH test, the pH of a sample 

is adjusted within a physiologically tolerable range 

(e.g., pH 6.0, 7.0, and 8.0) before toxicity testing. 

In many instances, the unionized form of a 

toxicant is able to cross biological membranes 

more readily than the ionized form and thus is 

more toxic. This test is designed primarily for 
Sediment Clnss~htk Methods Compendium 

F- 1 

( TOXIC AQUEOUS SAMPLE ( 

EDTA 

Chelation 

Aeration I C 18 Solid Phase 

1 Extraction 

1 pH AdjFstment 1 Graduated pH 

Test 

Acid PHI Base p”6 P”7 ~“8 

Figure 51. Overview of the Phase 1 Toxicity Characterization Process. 

The ambient pH of the sample is indicated as pH,. 

54 
S-Interstitial Water TIE Appmach 

ammonia, a relatively common toxicant whose 

toxicity is extremely pH-dependent (USEPA, 

198%). However, different pH values can strongly affect the toxicity of many common ionizable 

pesticides, and also may influence the bioavailability and toxicity of certain heavy metals and 

surfactants (Campbell and Stokes, 1985; Doe et 

al., 1988). 

Aeration tests are designed to determine 

whether toxicity is attributable to volatile, oxidizable, or sublatable compounds. Samples at pH, 

(ambient pH), pH 3, and pH 11 are sparged with 

air for 1 h, readjusted to pH, and tested for toxicity. The different pH values affect the ionization 

state of polar toxicants, thus making them more or 

less volatile, and also affect the redox potential of 

the system. If toxicity is reduced by air sparging 

at any of the pH values, the presence of volatile or 

oxidizable compounds may be suggested. To 

distinguish the former from the latter situation, 

further experiments are performed using nitrogen 

rather than air to sparge the samples. If toxicity 

remains the same, oxidizable materials are implicated; if toxicity is again reduced, volatile compounds are suspect. The pH at which toxicity is 

reduced is important. if nitrogen sparging decreases toxicity at pH,, neutral volatiles are present; if toxicity decreases at pH 11.0 or pH 3.0, 

basic and acidic volatiles, respectively, are implicated. An additional process through which 

aeration can remove sample toxicants is sublation, 

which is the movement of compounds through 

aqueous solutions at the surface of the air bubbles, 

often followed by deposition on the aeration glassware. Compounds that exhibit this behavior 

include resin acids and surfactants; in some instances it may be possible to implicate the presence of sublatable compounds by rinsing the 

aeration glassware with clean laboratory dilution 

water and testing this fraction (Ankley et aZ., 

1990b). 

Filtration provides information concerning the 

amount of toxicity associated with filterable 

components. In this test, samples at pH,, pH 3.0, 

and pH 11.0 are passed through 1-p glass fiber 

filters, readjusted to pH,, and tested for toxicity. 

Reductions in toxicity due to filtration could be 

related to factors such as decreased physical 

toxicity, rather than chemical toxicity (Chapman 

et al., 1987), or removal of particle-bound toxicants, which could be important, particularly if 

filter-feeding organisms such as cladocerans are 

the test species. 

Reversed-phase, solid-phase extraction (SPE) 

is designed to determine the extent of toxicity due 

to compounds that are relatively nonpolar at pH, 

pH 3.0, or pH 9.0. This test, in conjunction with 

associated Phase II analytical procedures, is an 

extremely powerful TIE tool. In this procedure, 

filtered sample aliquots at pH, pH 3.0, and pH 9.0 

are passed through small columns packed with an 

octadecyl (C,,) sorbent. At pH, the C,, sorbent 

will remove neutral compounds such as certain 

pesticides (Junk and Richard, 1988). By shifting 

ionization equilibria at the low and bigh pH 

values, the SPE column also can be used to 

extract organic acids and bases (Wells and Michael, 1987). During extraction, the resulting 

post-column effluent is collected and tested for 

toxicity to determine whether the manipulation 

removed toxicity and/or whether the capacity of 

the column was exceeded. Following this, the 

column is eluted with solvents, such as methanol, 

which then can be tested for recovery of toxicity. 

If sample toxicily is decreased and subsequently 

recovered in solvent elutions, a nonpolar toxicant 

would be suspected. 

The presence of toxicity due to cationic metais 

is tested through additions of ethylenediaminetetraacetic acid (EDTA), a strong chelating agent that 

produces nontoxic complexes with many metals. 

As with SPE chromatography, the specificity of 

the EDTA test for a class of ubiquitous toxicants 

makes it a powerful TIE tool. Cations chelated by 

EDTA include certain forms of aluminum, barium, 

cadmium, cobalt, copper, iron, lead, manganese, 

nickel, strontium, and zinc (Stumm and Morgan, 

1981). EDTA does not complex anionic forms of 

metals, and only weakly chelates certain cationic 

metals, such as silver, chromium, and thallium 

(Stumm and Morgan, 1931). 

The oxidant reduction test is designed to 

determine the degree of toxicity associated with 

chemicals reduced, or in some instances chelated, 

by sodium thiosulfate. The toxicity of oxidants 

such as chlorine, bromine, iodine, and manganous 
Sediment ClasGjiuatM1 Methods Compendium 

ions is neutralized by sodium thiosulfate, and 

metals such as copper, cadmium, and silver are 

chelated and rendered biologically unavailable 

(Hockett and Mount, 1990). Because sodium 

thiosulfate, like EDTA, has low toxicity to most 

aquatic organisms, a relatively wide range of 

concentrations can be tested. 

Phase II: Toxlunt Identification-Initial laboratory work in Phase II focuses on isolation of 

the toxicants using chemical fractionation techniques with toxicity tracking. The ideal isolation 

process would create a subsample that contains 

one chemical, the toxicant. Depending on the 

nature of the toxicants, wide differences in the 

techniques, as well as in the amount of effort 

required for fractionation, will occur. 

In general, after fractionation, instrumental 

analyses are performed on the toxic subsamples, 

and lists of identified chemicals are assembled 

for each subsample. For each chemical in a list, 

an LC, value is obtained, usually from the 

literature or occasionally from structure activity 

models (Institute for Biological and Chemical 

Process Analyses, 1986). By comparing conceotrations of the identified chemicals to their IX, 

values, a list of suspect toxicants is made. This 

list is then refined by actually determining LCW 

values for the suspects using the TIE test species. 

If only one toxicant is present, it should be easily 

identified. Fdr samples with multiple toxicants, 

identification becomes significantIy more protraded since interactions among toxicants may 

need to be examined. If none of the suspected 

toxicants appears to explain the toxicity, the true 

toxicants were probably not detected during 

instrumental analysis. Usually, additional separation and associated concentration steps are required to increase the analytical sensitivity for 

toxicant identification. 

The information obtained in Phase I provides 

the analytical roadmarks for performing the 

fractionation and identification tasks in Phase II. 

To illustrate the relationship between Phase I 

data and the analytical approaches employed in 

Phase II, results for two typical Phase I TIE 

evaluations are presented in Table 5-l. The 

Phase II methods and approaches appropriate for 

these examples are discussed below. 

In the first sample in Table 5-1, SPE reduced 

toxicity. In Phase II, the SPE column is eluted 

with graded, increasingly nonpolar methanol/- 

water solutions, and toxicity testing is performed 

on each fraction (Burkhard et al., 1990). Although solvents other than methanol are routinely 

used in analytical work with q, chromatography 

columns, the low toxicity of methanol to aquatic 

organisms (e.g., Lc, 21.5 percent for dadocelans) makes it a solvent of choice for toxicity 

tracking in the fractions. If no toxicity occurs in 

the fractions, the toxicants have been lost and 

further characterization (Phase 1) work is required. If toxicity occurs in the fractions, Phase 

II methods feature concentralion of the toxic 

methanoltiater fractions; high performance 

liquid chromatography fractionation of the coocentrate (again with a CJmethanoVwater solvent 

system), with concurrent toxicity testing of the 

fractions; and, ultimately, identification of suspected toxicants in the toxic fractions via gas 

chromatography/mass spectroscopy. For pore 

water TIE, toxicity caused by high log L, oonpolar organics is often not elutable with methanol. In these cases, it is useful to elute the SPE 

column with a less polar solvent (i.e., methylene 

chloride) (Schubauer-Berigan and Ankley, 1991). 

In the second sample, both EDTA additions 

and SPE reduced toxicity. The reduction of 

toxicity by EDTA strongly suggests the presence 

of toxic levels of cationic metals. Thus, Phase II 

procedures would include both metal analyses 

and the concentration, fractionation, and identification techniques described for nonpolar organics in the first example. If analyses identify 

specific metals at concentrations high enough to 

cause toxicity, various mass balance procedures 

~8~1 be used to define the portion of the sample 

toxicity due to the suspected metals and the 

portion of the toxicity due to the suspect noopolar compounds. 

Only a very small subset of possible Phase I 

results is shown in Table 5-1, particularly when 

one considers that three of the tests (aeration, 

filtra!ioo, SPE) are conducted at three different 

pH values. A complete discussion of the types 

of Phase I results that may be encountered and 
!54nterstitial Water TIE Appmach 

Table 5-l. Phase I Characterization Results and Suspect 

Toxicant Classification for Two Samples. 

Aeration NR 

SPE Rb 

Methanol fractions T 

Suspected toxicant classification Nonpolar organics 

NR 

R 

T 

Nonpolar organicwheavy 

IlWtdS 

SJR = No redudion in toxicity 

‘R = Reduction in toxicity. 

7 = Toxicity recovered. 

subsequent Phase II strategies that could be 

implemented is beyond the scope of this review. 

Phase III: Toxicant Confirmation-After Phase 

II identification procedures implicate suspected 

toxicants, Phase III is initiated to confirm that the 

suspects are indeed the true toxicants. Confirmation is perhaps the most critical step of the TIE 

because procedures used in Phases I and II may 

create artifacts that could lead to erroneous conclusions about the toxicants. Furthermore, there is 

a possibility that substances causing toxicity are 

different from sample to sample within a supposedly homogeneous geographic region. Phase III 

enables both situations to be addressed. The tools 

used in Phase IIl include correlation, relative 

species sensitivity, observation of symptoms, 

spiking, and mass balance techniques. In most 

cases, no single Phase III test is adequate to confirm suspects as the true toxicants; it is necessary 

to use multiple confirmation procedures. 

In the correlation, approach, observed toxicity 

is regressed against expected toxicity due to 

measured concentrations of the suspected toxicants 

in samples collected over time or from several 

sites within a location. For the correlation approach to succeed, temporal or spatial variation 

has to be wide enough to provide a range of 

values adequate for meaningful analyses. To use 

the correlation approach effectively when there are 

multiple suspect toxicants, it is necessary to 

generate data concerning the additive, antagonistic, 

and synergistic effects of the toxicants in ratios 

similar to those found in the samples. These data 

also are needed for the spiking and mass balance 

techniques described below. 

The relative sensitivity of different test species 

can be used to evaluate suspected toxicants. If 

two or more species exhibit markedly different 

sensitivities to a suspected toxicant in standard 

reference tests, and the same patterns in sensitivity 

are seen with the toxic pore water sample, this 

5-7 
Sehunt ClussifiAion Mt&ods Compendium 

provides evidence for the validity of the suspect 

being the true toxic-ant. 

hother Phase fI1 procedure is observation of 

symptoms (e.g., time to mortality) in poisoned 

animals. Although this approach does not necessarily provide support for a given suspecl, it can 

be used to provide evidence against a suspected 

toxicant. If the symptoms observed in a standard 

reference test with a suspected toxicant differ 

greatly from those observed with the pore water 

sample (which contains similar concentrations of 

the suspected toxicant), this is strong evidence for 

a misidentification. 

Confirmatory evidence also can be obtained 

by spiking samples with the suspect toxicants. 

While the results may not be conclusive, an 

increase in toxicity by the same proportion as the 

increase in concentration of the suspect toxicant in 

the sample suggests that the suspect is correct. To 

obtain a proportional increase in toxicity from the 

addition of a suspect toxicant when in fact it is 

not the true toxicant, both the true and suspect 

toxicants would have to have very similar toxicity 

levels and their effects would also have to be 

additive. 

Mass balance calculations a be used as 

confirmation steps when toxicity can be at least 

partially removed from the pore water sample, and 

subsequently recovered. This approach can be 

useful in instances when SPE removes toxicity. 

T%e methanol fractions eluted from the SPE 

column are evaluated individually for toxicity; 

these toxicities are summed and then compared to 

the total amount of toxicity lost from the sample. 

Other techniques, including alteration of water 

quality characteristics (e.g., pH, salinity) in a 

manner designed to affect the toxicity of specific 

compounds, and analysis of body burdens of 

suspected toxicants in exposed animals, also can 

be useful confirmation steps. 

5.2.1.23 Types of Data Required 

In addition to the routine measurements desaiied above, biological response data, either acute 

or chronic, will be obtained. Specific data collected 

will depend on the choice of test organism and 

endpoints. If the TIE process is initiated, the 

researcher will fit obtain data amcerning the 

physicakhemical charackristics of the toxicants in 

the pore water, followed by actual identification of 

toxic compounds, and standard determination of 

their concentrations in the toxic samples (see Se.0 

tion 5.2.12.2 above). 

5112.4 Necessary Harti and Skills 

Pore water preparation and toxicity test proce 

dures are fairly straightforward and require cornmonIy available equipment and facilities. Many of 

the TIE procedures also require only routine fadlities. However, certain TIE techniques require some 

degree of advanced analytical capability (e.g., 

atomic absorption spectroscopy, gas rhromatographyhass spedroscopy). Similarly, although many 

of the routine toxicity tests require relatively little 

training, certain of the TIE procedures, in particular 

some of the chemical analyses, require advanced 

technical expertise and experience. 

52.13 Adequacy of Document&m 

‘Ihe theoretical basis for using pore water to 

assess toxicity appears to be scientifically sound, 

and pore water has been used for sediment toxicity 

evaluation (Adams et d., 1985; Swartz et al., 1985, 

1988,199@ Knezovich and Harrison, 1988; Connell 

et al., 1988; Giesy et ol., 1988; USEP& 198%; 

AnkIey et al, 199Oa, 1991a, 1991b; Hake et d, 

199Q Schubauer-Berigan and Ankley, 1991). 

Toxicity tests that can be used are in many instances 

well-documented, standard procedures (U.S. EPA, 

198Sa; 1985b). The TIE techniques involved, 

including those specifically for sediments, have been 

documented (USEPA, 1988,1989b, 1989c, 199la, 

1991b). Also, sediment TIES with pore water have 

been successfully demonstrated (AnkIey et al., 

199Oa, 1991b; Schubauer-Berigan et al., 1990, 

Schubauer-Berigan and Ankley, 1991). 

5.2.2 Applicability of Mctbod to Human 

Health, Aquatic IXe, or WildIKe 

ROtE!CtiOtl 

ilk method can be used to predict acute and 

chronic (i.e., grouti or reproductive) effects of toxic 

5-8 
54ntcrsCitid Wafer TIE Avach 

sediment on aquatic organisms and can identify 

toxicants responsible for observed effects. The data 

generated thus can be used to design sediment 

remediation programs that would have an optimal 

likelihood of success. These procedures are not 

suitable, however, for evaluating human health 

effects or protecting wildlife, and they cannot be 

used to address bioconcentratable toxicants. 

5.23 Ability of Method to Generate 

Numerical Criteria for Specific 

Chemicals 

Pore water toxicity assessment, in conjunction 

with successful TIE procedures, can be used to 

generate numerical criteria for toxic compounds in 

sediment pore water because the toxicants are 

actually identified. However, it must be established that compounds identified as being toxic to 

test organisms in the laboratory are the same 

compounds (both in form and concentration) 

responsible for toxicity to organisms in field 

situations. This relationship can be evaluated both 

through biosurveys (possibly in conjunction with 

analysis of contaminant residues in organisms 

collected from the field), and laboratory toxicity 

tests in which benthic organisms perceived to be 

affected in contaminated sediments in situ are 

exposed to toxicants identified in the pore water. 

Both types of data also would be required for any 

sediment classification method based on toxicity 

or chemical analyses. 

53 USEFULNESS 

53.1 Environmental ApplicabiliCy 

5.3.1.1 Suitability for Diflerent Sediment Types 

The pore water toxicity assessment approach 

is suitable for any sediment from which adequate 

quantities of pore water can be isolated. In typical 

sediments, 20-50 percent of the volume of the 

bulk sediment sample is pore water. For a complete Phase I characterization with a test species of 

relatively small body size (e.g., cladocerans, larval 

fishes), approximately 15 L of pore water is required. This translates into a bulk sediment 

requirement of 3-8 L Bulk sediment volumes 

needed for Phase II identification will, of course, 

be dependent on the toxicants present in the pore 

water, but typical volumes required would be 

expectedtorangefromlto2OL 

5.3.1.2 Suitability for Difirent Chemicals or 

This approach appears to be suitable for 

various nonpolar organics, cationic metals, and 

ammonia (Adams et al., 1985; Swartz et d., 1985, 

1988, 1990, Knezovich and Harrison, 1988; 

Connell et uZ., 1988; USEPA, 1989a; Ankley et 

al., 199Oa, 1991b; DiToro d al., 1990). The 

applicability of the approach to toxicants such as 

polar organics or extremely lipophilic compounds 

has yet to be established. Also, the TIE procedure-s enable the evaluation of interactive (additive, synergistic, antagonistic) effects among 

various toxicants present in pore water samples. 

5.3.1.3 Suitability for Predicting Effects on 

Diferent Organisms 

If the TIE procedures successfully identify 

specific toxicants responsible for sediment toxicity, the impacts of these toxicants on various 

species of concern can be easily predicted, provided that there are data concerning the toxicity of 

the identilied compounds to these species. Although toxicity data may not be available for 

certain benthic species, once suspect toxicants are 

identified, it would be possible to generate toxicity 

data for specific species of concern. 

5.3.1.4 Suitability for In-Place Pollutant Control 

The pore water toxicity assessment method 

and associated TIE procedures could prove to be 

a powerful tool for in-place pollutant control. Because sediment toxicants are actually identified, it 

is possible to design remediation plans for toxicants from point sources or controllable nonpoint 

sources, and to routinely monitor the success of 

these plans through continued assessm ent of pore 

water for toxicity and specific chemical toxicants. 

s-9 
Sediment Class$c&m Methods ComFdium 

5.3.1.5 Suitability for Source Control 

Because the potential exists for identifying 

specific sediment toxicants, this method is ideal 

for point source control, as well as controflabte 

nonpoint source inputs. 

5.3.1.6 Suitability for Dispsal Applications 

As stated above, because specific sediment 

toxicants can be identified, it would be possible to 

identify potential hazards of contaminated sediments to aquatic organisms before disposal operations, such as those associated with dredging 

(Ankley et al., 1991c). 

53.2 General Advantages and Umitatioas 

5.3.2.1 Ease of Use 

Pore water preparation, routine chemical 

analyses, toxicity tests, and certain of the TIE 

procedures are reasonably straightforward and 

require relatively little technical expertise or 

extensive laboratory facilities. Because it is 

possible to work with aqueous samples, many of 

the standard toxicity tests developed for toxicity 

assessment of surface waters and effluents can be 

used, in addition to tests with various benthic 

species (e.g., USEPA, 1985a, 1985b). However, 

interpretation of results of certain of the TIE 

procedures, as well as analytical support for the 

TIE work, requires advanced training and experience. Also, several TIE analyses require highly 

sensitive analytical instrumentation for procedures, 

such as atomic absorption spectroscopy and gas 

chromatography/mass spectroscopy. 

5.3.2.2 Relative Cost 

Cost of the actual toxicity rest procedures is 

relatively low. Cost of the TIE procedures will 

vary depending on the nature of the toxic compounds; certain toxicants (e.g., pesticides) are 

more costly to identify and quantify than others 

(e.g., ammonia). Also, identification and determination of the effects of multiple toxicants in 

samples costs more than the identification of 

single toxicants. Thus, cost analysis for the TIE 

portion of the toxicity assessment is case-specific. 

5.3.2.3 Tendency to Be Conservative 

Depending on the species used and the endpoint evaluated, pore water toxicity tests can be as 

conservative as desired. However, acute pore 

water toxicity tests described for sediment TfE are 

not.meant to represent chronic or bioaccumulation 

endpoints. 

5.3.2.4 Level of Acceptance 

The theoretical basis of pore water toxicity 

assessment is sound (Adams et al., 1985; Swartz 

et al. 1985, 1988, 1990; Knezovich and Harrison, 

1988; Connell et al., 1988; USEPA, 1989a, 

DiToro et ol., 1990, Ankley et al., 1991a). The 

most important advantage of using pore water as 

a sediment test fraction, however, is the fact that 

it enables the application of recently developed 

TIE procedures for the identification of toxic 

compounds in aqueous samples containing complex mixtures of chemicals (USEPA, 1988,198%, 

1989c, 1991a, 1991b). TIE procedures have 

proven to be extremely powerful tools for work 

with both complex effluents and sediment pore 

water (Ankley et al., 199Oa, 1991b; Kuehl et al., 

1991; Amato et al., 1991; Norberg-King et al., 

1991; Schubauer-Berigan and Ankley, 1991; 

Ankley and Burkhard, 1992). The ability to 

identify specific compounds responsible for the 

toxicity of contaminated sediments clearly could 

be critical to the success of remediation. 

5.3.2.5 Ability to Be Implemented by 

L&oratories with Qpical Equipment 

and Handling Facilities 

Pore water preparation, toxicity test pmdures, and certain of the TIE methods are easily 

implemented by laboratories with typical equip 

ment and a moderate degree of expertise. Interpretation of some TIE results requires additional 

technical training and experience, and certain of 

the analytical procedures associated with TIE 

5-10 
S-hzterstifial Water TIE Approach 

work require both specialized training and analytical instrumentation. 

5.3.2.6 Level of Effort Required to Generate 

Results 

This procedure consists of field sampling, 

preparation of pore water, toxicity tests, and 

various TIE procedures. Depending on the 

results of the TIE work, the level of effort expended to obtain potentially important data can 

be relatively small. 

5.3.2.7 Degree to Which Results Lend 

Themselves to Interpretation 

Biological responses (i.e., toxicity) can be 

easily interpreted, and when properly performed, 

the results of the TIE procedures can be straightforward and easily interpreted; however, this is 

dependent on the complexity of the sample and 

the number of compounds contribuling to sample 

toxicity. 

5.3.2.8 Degree of Environmental Applicability 

Pore water toxicity assessment and TIE 

procedures are applicable to virtually all environmental conditions and sediment types. 

Moreover, a wide variety of test organisms can 

be evaluated with this approach. However, 

although data indicate that the toxicity and/or 

bioaccumulation of a variety of contaminants are 

correlated with their pore water concentrations, 

there is no guarantee that this relationship exists 

for all types of contaminants. For example, a 

potentially important route of exposure for 

highly lipophilic compounds is thought to be via 

ingestion of contaminated particles. This route 

is not addressed using pore water exposures. 

Finally, existing TIE procedures are applicable 

for acutely toxic samples, and thus generally 

would not be useful for identifying chronically 

toxic sediment contaminants. 

5.3.2.9 Degree ofAccuracy and Precision 

Because the procedures consist of laboratory-controlled experiments, results obtained are 

statistically accurate and precise. 

5.4 STATUS 

5.4.1 Extent of Use 

Various toxicity tests have been widely applied to the evaluation of both freshwater and 

marine sediments, and pore water is merely one of 

the possible fractions that can be tested. Theoretically, pore water appears to be appropriate for 

sediment toxicity assessment and there have been 

many examples of its use for this purpose (Adams 

et al., 1985; Swartz et al., 1985, 1988, 1990; 

Giesy et al., 1988; Knezovich and Harrison, 1988, 

Connell et al., 1988; USEPA, 1989a; Ankley, 

199Oa, 1991a, 1991b; DiToro et al., 1990, Hake 

et al., 1990; Schubauer-Berigan and Ankley, 

1991). The TIE procedures (USEPA, 1988, 

1989b, 1989c, 1991a, 1991b) although developed 

only relatively recently, already are widely used in 

both research and regulatory programs. 

5.4.2 Extent to Which Approach Has Beea 

F&M-V&i&d 

Because the procedure is relatively new, there 

has been little field validation. ‘Ihis area requires 

research, not only for the pore water TIE methods 

desuibed herein, but for virtually any other sediment 

method involving toxicity tests or chemical analyses. 

SA.3 Reasms for Ihlted Use 

Various sediment toxicity tests have been widely 

used; however, relatively few studies have evaluated 

pore water toxicity. nis is primarily because the 

theoretical basis for using pae water has only 

recently been uitically evaluated. For this reaso& 

there are no standard methods for pore water prcparation. Systematic TIE procedures for toxic aqueous 

samples have only recently been developed and thus 

have not yet been widely applied to the area of 

sediment toxicity assessment. Because avrent TIE 

pmcedures cannot be used with bulk sediment samples, pore water appears to be the best fraction with 
Sediment Classifktion Methods Compendium 

which to attempt to identify specific sediment 

contaminants responsible for acute toxicity. 

5.4.4 Outlook for Future Use and Amount of 

Development Yet Needed 

The outlook for this approach is extremely 

promising because it is the only method currently 

available that enables the identification of specific 

compounds responsible for sediment toxicity with 

some degree of certainty. This information could 

be critical to the success of remediation. However, as with all of the existing sediment methods, 

further development is needed, particularly in the 

following areas: 

. The development of standard and scientifically sound techniques for pore water 

isolation; 

m Further characterization of relationships 

between sediment toxicity h sihc and the 

toxicity of sediment pore water in the 

laboratory for different classes of campounds; and 

m The development of TIE procedures to 

identify chronically toxic compounds in 

aqueous samples. 

Research in all these areas is ongoing at ERL 

Duluth. 

For more information please contact: 

Gerald Ankley and Nelson Thomas 

U.S. Environmental Protection Agency 

Environmental Research Laboratory 

6201 Congdon Boulevard 

Duluth, MN 55804 

(218) 7205603 

Mary K. Schubauer-Berigart 

AScI Corporation 

6201 Congdon Boulevard 

Duluth, MN 55804 

(218) 720-5619 

5.S REFERENCES 

Adams, W.J., R.A. Kimerle, and R.G. Mosher. 

1985. Aquatic safety assessment of chemicals 

sorbed to sediments. pp. 429453. In: Aquatic 

Toxicology and Hazard Assessment: Seventh 

Symposium. R.D. Cardwell, R. Purdy, and R.C. 

Bahner (eds.). ASTM STP 854. American Stiety for Testing and Materials, Philadeiphia, PA. 

Amato, J.R., D.I. Mount, EJ. Durhan, M.T. 

Lukasewycz, G.T. Ankley, and E.D. Robert. 

1991. An example of the identification of 

diazinon as a primary toxicant in an effluent. 

Environ. Toxicol. Chem. In press. 

Ankley, G.T. and LP. Burkhard. 1992. Identification of surfactants as toxicants in a primary 

effluent. Environ. Toxicol. Chem. Submitted. 

hkley, G.T., A. Katko, and J.W. Arthur. 199Oa. 

Identification of ammonia as an important sediment-associated toxicant in the lower Fox River 

and Green Bay, Wisconsin. Environ. Toxicol. 

Chem. 9:313-322 

Ankley, G.T., M.T. Lukasewycz, G.S. Peterson, 

and DA. Jenson. 199Ob. Behavior of surfactants in toxicily identification evaluations. 

Chemosphere. 21:3-12. 

Ankley, G.T., M.K. Schubauer-Berigan, and J.R. 

Dierkes. 1991a. Predicting the toxicity ol bulk 

sediments to aquatic organisms with aqueous 

test fractions: Pore water versus elutriate. 

Environ. Toxicol. Chem. In press. 

Ankley, G.T., G.L Phipps, P.A. Kosian, DJ. Hansen, J.D. Mahony, A.M. Cotter, E.N. Leonard, 

J.R. Dierkes, DA. Benoit, and V.R. Mattson. 

1991b. Acid volatile sulfide as a factor mediating cadmium and nickel bioavailability in 

contaminated sediments. Environ. Toxicol. 

Cbem. In press. 

Ankley, G.T., M.K Schubauer-Berigan, and RA. 

Hoke. 1991~. Use of toxicity identification 

evaluation techniques to identify dredged material disposal options: A proposed approach. 

Environ. Management. In press. 

Burkhard, LP., and G.T. Ankley. 1989. 

NETAC’s toxicity-based approach to identify 

toxicants. Environ. Sci. Technol. 23:1438- 

1443. 

s-12 
CInterstitial Water TZE Appmach 

Burkhard, LP., E.J. Durban, and M.T. 

Lukasewycz. 1990. Identification of nonpolar 

toxicants in effluent using toxicity-based fractionation with gas chromatographybass spectrometry. Anal. Chem. 63:277-283. 

Campbell, P.G.C., and P.M. Stokes. 1985. 

Acidification and toxicity of metals to aquatic 

biota. Can. J. Fish. Aq. Sci. 42:2034-2049. 

aapman, P.M., I.D. Popham, J. Griffin, D. 

Leslie, and J. Michaelson. 1987. Differentiation of physical from chemical toxicity in solid 

waste fish bioassays. Water Air Soil Pollut. 

33:295-308. 

Connell, D.W., M. Bowman, and D.W. Hawker. 

1988. Bioconcentration of chlorinated hydrocarbons from sediment by oligochaetes. Ecotoxicol. Environ. Safety 16:293-302. 

DiToro, D-M., J.D. Mahony, D.J. Hansen, K.J. 

Scott, M.B. Hicks, S.M. Mays, and M.S. Redmond. 1990. Toxicity of cadmium in sediments: the role of acid volatile sulfide. 

Environ. Toxicol. Chem. 9: 1489-1504. 

Doe, KG. W.R. Ernst, W.R. Parker, GRJ. Julien, 

and PA. Hennigar. 1988. Influence of pH on 

the acute lethality of fenitrothion, 24-D and 

aminocarb and some pH-altered sublethal effects 

of aminocarb on rainbow trout (&r/mo gairdnerd). Can. J. Fish. Aq. Sci. 45287-293. 

Giesy, J.P., R.L. Graney, J.L. Newsted, C.J. 

Rosiu, A. Benda, R.G. Kreis, and F.J. Horvath. 

1988. Comparison of three sediment bioassay 

methods using Detroit River sediments. Environ. Toxicol. C&em. 7483-498. 

Hackett, J.R., and D.R. Mount. 1990. Use of 

metal chelating agents to differentiate among 

sources of toxicity. Eleventh Annual Meeting 

of the Society of Environmental Toxicology and 

Chemistry, Abstract, p. 162. 

Hoke, RA., J.P. Giesy, G.T. Ankely, J.L Newsted, and J.R. Adams. 1990. Toxicity of 

sediments from western Lake Erie and Maumee 

River at Toledo, Ohio, 1987: Implication for 

current dredged material disposal practices. J. 

Great Lakes Res. 16:457-470. 

Institute for Biological and Chemical Process Analyses. 1986. User manual for QSAR system. 

Montana State University, Boxernan, MT. 

Junk, GA., and J.J. Richard. 1988. Organics in 

water: Solid phase extraction on a small scale. 

Anal. C%em. 60~451-454. 

Knezovich, J-P., and F.L Harrison. 1988. ‘The 

bioavailability of sediment sorbed chlorobenzenes to larvae of the midge Chiromw 

&corus. Ecotoxicol . Environ. Safety 

15226241. 

Knezovich, J.P., F.L Henderson, and R.G. Wilhelm. 1987. The bioavailability of sediment-- 

sorbed organic chemicals: A review. Water Air 

Soil Pollut. 32233245. 

Kuehl, D.W., G.T. Ankley, LP. Burkhard, and 

DA. Jensen. 1990. Bioassay directed characterization of the acute toxicity of a creosote 

leachate. Hazardous Waste Hazardous Mater. 

7283-291. 

Norberg-King, TJ., E.J. Durhan, G.T. Ankley, and 

E. Robert. 1991. Application of toxicity identification evaluation procedures to the ambient 

waters of the Colusa Basin Drain. Environ. 

Tox. and C&em. In press. 

Schubauer-Berigan, M.K., J.R. Die&es, and G.T. 

Ankley. 1990. Toxicity identification evaluations of contaminated sediments in the Buffalo 

River, NY and Saginaw River, MI. National 

Effluent Toxicity Assessment Center Rep. No. 

20-90. Environmental Research Laboratory, 

Duluth, MN. 

Schubauer-Berigan, M.K., and G.T. Ankley. 

1991. The contribution of ammonia, metals, 

and nonpolar organic compounds to the toxicity 

of sediment interstitial water from an Illinois 

River tributary. Environ. Toxicol. Chem. In 

press. 

Shults, D.W., LM. Smith, S.P. Ferraro, F.A. Roberts, and C.K. Poindexter. 1991. A comparison 

of methods for measuring trace organic compounds and metals in interstitial water. Water 

Res. In press. 

Sly, P.G. 1988. Interstitial water quality of lake 

trout spawning habitat. J. Great Lakes Res. 

14:301-315. 

Stumm, W., and J.J. Morgan. 1981. Aquatic 

chemistry - An introduction emphasizing chemical equilibria in natural waters. John Wiley and 

Sons, New York. 583 pp. 

5-13 
Scdimrn t Class$cn tion Methods Compendium 

Swartz, R.C, G.R. Ditswortb, D.W. Schults, and 

J.O. Lamberson. 1985. Sediment toxicity to a 

marine infaunal amphipod: Cadmium and its 

interaction witb sewage sludge. Mar. Environ. 

Res. 18: 133-153. 

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. Environ. 

Toxicol. Cbem. 7:1013-1020. 

Swartx, R.C., P.F. Kemp, D.W. Scbults, G.R. 

Ditsworth, and R.J. Ozretich. 1989. Acute 

toxicity of sediment from Eagle Harbor, Washington, to the infaunal amphipod Rhepdlrynius 

abronius. Environ. Toxicol. Qem. 8:215-222. 

Swartz, R.C., D.W. Schults, T.H. Dewitt, G.R. 

Ditsworth, and J.O. Lunberson. 1990. Toxi-city of fluoranthene in sediment to marine amphipods: A test of the equilibrium partitioning 

approach to sediment quality criteria. Environ. 

Toxicol. Chem. 9:1071-1080. 

USEPA. 1985a. Methods for measuring the 

acute toxicity of effluents to freshwater and 

marine organisms. EP~600/485-013. U.S. 

Environmental Protection Agency, Cincinnati, 

OH. 

USEPA. 1985b. Short-term methods for estimating the chronic toxicity of effluents and receiving waters to freshwater organisms. 

EPA/600/4-85-014. U.S. Environmental Protection Agency, Cincinnati, OH. 

USEPA. 198%. Ambient water quality criteria 

for ammonia - 1984. EPA/440/5-85001. U.S. 

Environmental Protection Agency, Duluth, MN. 

USEPA. 1988. Methods for aquatic toxicity 

identification evaiuations: Pbase I toxicity 

characterization procedures. EPA/600-3-88/034. 

U.S. Environmental Protection Agency, Duluth, 

MN. 

USEPA. 1989a. Equilibrium partitioning 

approach to generating sediment quality criteria. 

EPA/440/5-89/002. U.S. Environmental 

Rote&on Agency, Washington, DC. 

USEPA. 1989b. Methods for aquatic toxicity 

identification evaluations: Phase II toxicity 

identification procedures. EPA/600-3-88/035. 

U.S. Environmental Protection Agency, Duluth, 

MN. 

USEPA. 1989c. Methods for aquatic toxicity 

identification evaluations: Phase III toxicity 

confirmation procedures. EPA/600-3-88/036. 

U.S. Environmental Protection Agency, Duluth, 

MN. 

USEPA. 1991a. Methods for aquatic toxicity 

identification evaluations: Phase I toxicity 

characterization procedures. Second edition. 

EPA-600/6-91/003. Environmental Research 

Laboratory, Duluth, MN. 

USEPA. 1991b. Methods for sediment toxicity 

identification evaluations. National EffIuent 

Toxicity Assessment Center Rep. No. 08-91. 

Environmental Research Laboratory, Duluth, 

MN. 

Wells, M.J.M., and J.L. Michael. 1987. 

Reversed-phase solid-phase extraction for aqueous environmental sample preparation in herbicide residue analysis. J. Chromatogr. Sci. 

25:345-50. 

5-14 
CHAPTER 6 

Equilibrium Partitioning Approach 

Christopher S. Zarba 

U.S. Environmental Protection Agency 

401 M Street, SW (WH-586), Washington, DC 20460 

(202) 260-1326 

The equilibrium partitioning (EqP) approach 

focuses on predicting the chemical interaction 

among sediments, interstitial water (i.e., the 

water between sediment particles), and contaminants. Based on correlations with toxicity, 

interstitial water concentrations of contaminants 

appear to be better predictors of biological 

effects than do bulk sediment concentrations. 

The EqP method for generating sediment quality 

criteria is based on predicted contaminant concentrations in interstitial water. Chemically 

contaminated sediments are expected to cause 

adverse biological effects if the predicted interstitial water concentration for a given contaminant exceeds the chronic water quality criterion 

for that contaminant. 

6.1 SPECIFIC APPLICATIONS 

Specific applications of EqP-based sediment 

quality criteria are under development. The 

primary use of EqP-based sediment criteria will 

be to identify and prevent risks associated with 

contaminants. Because the regulatory needs 

vary widely among and within U.S. EPA offices 

and programs, EqP-based sediment quality 

criteria will be used in a variety of ways. 

EqP-based numerical sediment quality 

criteria would likely be used directly to assess 

risk and would be applied in a tiered approach 

In tiered applications, concentrations of sediment 

contaminants that exceed sediment quality 

criteria would be considered as causing unacceptable impacts. Further testing may or may 

not be required, depending on site-specific and 

program-specific conditions. Sediment contaminants at concentrations less than the sediment 

criteria would not be of concern. However, 

sediments would not be considered safe in cases 

where they are suspected to contain other contaminants at concentrations above safe levels, 

but for which no sediment criteria exist. 

Synergistic, antagonistic, or additive effects 

of multiple contaminants in the sediments may 

also be of concern. Additional testing in other 

tiers of the evaluation approach, such as bioassays, could be required to determine whether 

the sediment is safe. It is likely that such 

testing would incorporate site-specific considerations. 

6.1.1 Current Use 

Specific regulatory uses of EqP-based sediment quality criteria are under development and 

will be articulated in the Contaminated Sediment 

Management Strategy. The Science Advisory 

Board (SAB) has completed the review of this 

approach for nonionic organic contaminants. 

Based on the findings of this review, the method 

will be used for developing national sediment 

quality criteria. (The first five sediment quality 

criteria will be proposed in the Federal Register 

shortly for public comment.) At the present 

time, the criteria are for the protection of benthic organisms. The methodology for developing sediment criteria for metal contaminants will 

be presented to the SAB for review in 1993. 

The range of potential applications of the EqP 

approach is large because the approach accounts 

for contaminant bioavailability and can be used 

to evaluate most sediments. 

Draft sediment criteria values have been 

developed for a variety of organic compounds 

using the EqP approach. In pilot studies at a 

variety of contaminated sediment sites at which 

site characterization and evaluation activities 

were undertaken, the draft criteria were used in 

the following ways: 
Sediment Classification Methods Compendium 

• Identify extent of contamination; 

• Assess the risks or potential risks associated with the sediment contamination; 

• Identify responsible parties and the need 

for source controls; and 

• Identify the environmental benefit associated with a variety of remedial options. 

In addition, a number of states have used draft 

EqP-based sediment criteria to evaluate the potential effects of sediment contaminants found in 

aquatic habitats. 

6.1.2 Potential Use 

Potential applications of the EqP approach 

include a variety of ongoing activities conducted 

by the U.S. EPA. EqP-based sediment quality 

criteria could play a major role in the identification, monitoring, and cleanup of contaminated 

sediment sites on a national basis. This is true, in 

part, because EqP-based SQC establish a direct 

cause-and-effect relationship between a contaminant concentration and biological impacts. They 

could also be used to ensure that uncontaminated 

sites remain uncontaminated. In some cases, such 

sediment criteria alone will be sufficient to identify and establish cleanup levels for contaminated 

sediments. In other cases, it will be necessary to 

supplement the sediment criteria with biological 

sampling, testing, or other types of analysis before 

a decision can be made. 

EqP-based sediment criteria will be particularly valuable at sites where sediment contaminant 

concentrations are gradually increasing. In such 

cases, criteria will permit an assessment of the 

extent to which unacceptable contaminant concentrations are being approached or have been exceeded. Comparisons of field measurements to 

sediment criteria will be a reliable method for 

providing an early warning of a potential problem. 

Such an early warning would provide an opportunity to take corrective action before adverse 

impacts occur. 

Although sediment criteria developed using 

the EqP approach are similar in many ways to 

existing water quality criteria, their applications 

may differ substantially. In most cases, contaminants in the water column need only be controlled 

at the source to eliminate unacceptable adverse 

impacts. In contrast, contaminated sediments 

often have been in place for quite some time, and 

controlling the source of that pollution (if the 

source still exists) will not be sufficient to alleviate the problem. Safe removal, treatment, or 

disposal of contaminated sediments can also be 

difficult and expensive. For this reason, it is 

anticipated that EqP-based sediment criteria will 

rarely be used as mandatory cleanup levels. 

Rather, they will likely be used to predict or 

identify the degree and spatial extent of problems 

associated with contaminated areas, and thereby 

facilitate regulatory decisions. 

6.2 DESCRIPTION 

6.2.1 Description of Method 

Concentrations of contaminants in the interstitial water correlate very closely with toxicity, 

whereas concentrations of contaminants bound to 

the sediment particles do not. The EqP method 

for generating sediment criteria involves predicting 

contaminant concentrations in the interstitial water 

and comparing those concentrations to quality 

criteria. If the predicted sediment interstitial water 

concentration for a given contaminant exceeds its 

respective chronic water quality criterion, then the 

sediment would be expected to cause adverse 

effects. 

The processes that govern the partitioning of 

chemical contaminants among sediments, interstitial water, and biota are better understood for 

some kinds of chemicals than for others. Concentrations of sulfides and organic carbon have 

been identified as primary factors that control 

phase associations, and therefore bioavailability, 

of trace metals in sediments. However, models 

that can use these factors to predict research are 

not fully developed. Mechanisms that control 

the partitioning of polar organic compounds are 

6-2 
6-EqP Approach 

also poorly understood. Polar organic contaminants, however, are not generally considered to 

be a significant problem in sediments. Partitioning of nonionic organic compounds between 

sediments and interstitial water is highly correlated with the organic carbon content of sediments. Also, the toxicity of nonionic organic 

contaminants in sediments is highly dependent 

on their interstitial water concentrations. Consequently, to date, the EqP approach is well 

developed for nonionic organic contaminants 

-.nd is in the process of development for trace 

metals. 

Interstitial water concentrations can be 

calculated using partition coefficients for specific nonionic organic chemicals and criteria continuous concentrations from WQC documents. 

The sediment quality criterion for a specific 

chemical is defined as the solid phase concentration that will result in an uncomplexed interstitial water concentration equal to the chronic 

water quality criterion for that chemical. The 

rationale for using water quality criteria as the 

effect concentrations for bentbic organisms is 

that the sensitivity range for bentbic organisms 

appears to be similar to the sensitivity range for 

water column organisms. Moreover, partition 

coefficients for a wide variety of contaminants 

are available. 

The calculation procedure for nonionic 

organic contaminants is as follows: 

where: 

cWQC = 

rSQC = 

16 = 

Criterion continuous concentration 

Sediment quality criterion 

@g/kg sediment) 

Partition coefficient for the 

chemical (L&g sediment) 

between sediment and water. 

Although the method for developing sediment 

criteria for nonionic organic contaminants has 

been identified, continuous refinement of the 

methodology is expected. 

6.2.1.1 Objectives ad Assumptions 

Three principal assumptions underlie use of 

the EqP-based approach to establisb sediment 

quality criteria: 

For sediment-dwelling organisms, the 

uncomplexed interstitial water concentration of a chemical correlates with observed biological effects across sediment 

types, and the concentration at which 

effects are observed is the same as that 

observed in a water-only exposure. 

Partitioning models permit calculation of 

uncomplexed interstitial water concentrations of the chemical phases of sediments 

controlling availability. 

Benthic organisms exhibit a range of 

sensitivities to chemicals that is similar to 

the range of sensitivities exhibited by 

water column organisms. 

Data exist supporting each of these assumptions. 

6.2.1.2 Level 0fEfior-t 

6.2.1.2.1 Type of Sampling Required 

Sufficient sediment chemistry sampling is 

required to adequately characterize the area of 

concern. Total organic carbon concentrations are 

also needed, preferably for each sampling station. 

6.2.1.2.3 Types of Data Required 

Analyses are needed to determine the concentrations of the contaminants of concern in the 

sediment (bulk sediment analysis) and the concentrations of organic carbon in the sediment. 

6.2.1.2.4 Necessary Hardware and Skills 

The investigator must be able to design an 

appropriate sampling study, conduct bulk sediment 

analyses, operate a pocket calculator, and understand developed values and what they protect. 

6-3 
S&nent Classification Methods Compendium 

6.2.1.3 Adequacy of Documentation 

The method is very well documented (see 

Section 6.5). 

6.2.2 Appikability of Metbod to Human 

Health, Aquatic Ufe, or Wildlife 

Protection 

At the present time SQC do not address 

bioaccumulative impacts to aquatic life, wildlife, 

and human health. Efforts are under way to 

derive criteria protective of these endpoints. 

6.23 Ability of Metbod to Generate 

Numerical Criteria for Specifk 

Chemicals 

The EqP method generates numerical criteria 

for a number of nonionic organic chemicals. A 

methodology for developing sediment criteria for 

metal contaminants is being developed. Draft 

criteria to be proposed in the Federal Register 

were developed for endrin, phenanthrene, fiuoranthene, dieldrin, and acenaphthene. It is expected that three to five additional sediment criteria 

will be issued each subsequent year. 

Methods for developing sediment criteria for 

metal contaminants are under development and are 

expected to be reviewed by the SAB in 1993. 

63 USEFULNESS 

63.1 Environmental ApplicablIity 

One of the principal reasons for selecting the 

EqP approach is that it is applicable in a wide 

variety of aquatic systems, which is a prerequisite 

for the development of national sediment quality 

criteria. 

6.3.1.X Suitability for Different Sediment Types 

Although aspects of the EqP method are still 

under development, it is expected that sediment 

criteria for nonionic contaminants developed using 

this approach will be applicable to all types of 

sediments found in both freshwater and marine 

environments with organic carbon concentrations 

So.2 percent organic carbon. Additional work is 

needed to clarify the best use of the EqP approach 

for sediments with less than 0.2 percent organic 

carbon. 

6.3.2.2 Suitability for Difirent Chemicals or 

Classes of Chemicals 

The EqP method for developing sediment 

criteria has been modified for different types of 

contaminants. Nonionic, ionic, and metal contaminants all interact with sediment particles in 

different ways, and partitioning models have to be 

modified to account for these differences. The 

technical approach for developing sediment criteria for nonionic organic contaminants has been 

well developed and is under peer review. The 

technical approach for developing sediment ait&a for metal contaminants is under development 

and is expected to undergo peer review in 1993. 

Ionic contaminants are not believed to cause major 

problems in sediments, but work plans for sediment criteria development methods for these 

compounds have been written. 

6.3.1.3 Suitabili~ for Predicting Effects on 

Different Organisms 

As indicated above (see Section 6.2.1), the 

EqP approach is based on predicted interstitial 

water concentrations of nonionic organic contaminants, and comparisons of these concentrations with chronic water quality criteria. Typically, water quality criteria are based on toxicity 

information (e.g., median lethal or median effective concentrations) for a wide number of species 

and are set low enough to be protective of at least 

95 percent of the species tested. Consequently, 

exposure levels that are predicted using the EqP 

approach can be compared with a range of toxic 

effects values that are representative of the diffefent kinds of organisms on which water quality 

criteria are based. 

64 
6-EqP Approach 

6.3.1.4 Suitability for In-Place Pollutant 

Control 

The EqP method is suitable for in-place 

pollution control because it can be used to 

identify locations where concentrations of individual contaminants are causing adverse effects. 

Target cleanup levels can be identified, and the 

success of cleanup activities can be determined. 

6.3.1.5 Suitability for Source Control 

The EqP method is suitable for source 

control. This metbod predicts the concentration 

of a contaminant above which adverse impacts 

are likely. A direct measure of biological 

effects is not needed to identify safe levels. 

6.3.1.6 Suitability for Disposal Applications 

The EqP method is suitable for predicting 

the effects that contaminated sediments may 

have if moved to an aquatic site. It is not 

applicable to contaminated sediments that are 

disposed of at upland sites. 

6.3.2 General Advantages and Limitations 

The EqP approach offers the following 

advantages: 

n It is consistent with existing water quality criteria; 

n It establishes a cause-and-effect relationship; 

m It relates risks to specific substances, 

and it can be used to identify probable 

species at risk; 

n It is applicable across all types of sediments and in all types of aquatic environments, including lentic, lofic, marine, 

and estuarine environments; 

n Only site-specific chemistry data are needed; 

n 

m 

m 

m 

The E 

ways: 

l 

l 

l 

l 

Site-specific or station-specific sediment 

criteria can be calculated as soon as sediment chemistry data are available; 

It incorporates the large quantities of data 

that were used in the development of 

water quality criteria; 

It can be incorporated into existing regulatory mechanisms with little or no need 

for additional staffing or skills, 

The equilibrium partitioning theory on 

which it is based is well developed; 

It can be modified easily to accommodate 

site-specific circumstances; 

It can be used with additional development to identify risks to humans and 

wildlife that may occur as a result of 

bioaccumulation; and 

It identifies the degree of sediment contamination and permits an assessment of 

whether contaminant concentrations are 

approaching an effects level. 

qP approach is limited in the following 

Sediment criteria developed using this approach do not address possible synergistic, antagonistic, or additive effects of 

contaminants; 

Interim and draft sediment criteria presently exist for only 12 contaminants at 

this time; 

The technical approach for developing 

sediment criteria for metal contaminants is 

still under development; 

Sediment quality criteria for nonionic 

chemicals apply to sediments that have an 

organic carbon concentration a.2 percent; 

and 

6-5 
Sedimenf Classification Methods Compendium 

n Sufficient water-only toxicity data do not 

exist for all aMaminants of concern. 

6.3.2.1 Ease of Use 

The calculation of site-specific sediment criteria 

is relatively easy, provided that sediment chemistry 

data adequately characterizing the site, a partition 

coefficient, and water quality criteria protective of 

the desired organism are available. 

6.3.2.2 Relatix? Cost 

Because site-specific biological data are not 

needed, the costs associated with this method 

depend primarily on the cost of collecting sitespecific chemistry data. 

6.3.2.3 Tendency to Be Consemtive 

Sediment criteria are derived using the chronic 

water quality criteria as effect levels. Hence, the 

levels of protection afforded by sediment aiteria are 

similar to those of water quality criteria. In general, 

water quality criteria are deemed to be protective of 

95 percent of the organisms most of the time. Each 

SQC is bracketed with levels of uncertainty. 

6.3.2.4 Level of Acceptance 

The EqP approach and its use in deriving 

sediment quality criteria are the result of the efforts 

of many scientists who represent a variety of federal 

agencies, industries, environmental organizations, 

universilies, U.S. EPA laboratories, state agencies, 

and other institutions. These scientists were involved in the selection of the EqP approach for 

generating sediment criteria and have also played a 

role in development of the method. Papers that 

discuss various aspeas of this effort have been 

presented at scientific conferences. 

6.3.2.5 Ability to Be Implemented by Laboratories 

with Typical Equipment and Handling 

Facilities 

No special laboratory facilities or requirements 

are needed. Sediment chemistry analysis is all that 

is required. 

6.3.2.6 Level of Effort Required to Generate 

Results 

The necessary level of effort varies substantially from site to site and is dependent on many 

factors. Compared with other methods, the EqP 

method generates results quickly and more costeffectively. No site-specific biological data are required. 

6.3.2.7 Degree to Which Results Lend 

Themselves ‘to Interpretarion 

All sediment evaluation procedures require 

some level of interpretation. However, a sediment 

criterion that is bracketed with an appropriate 

degree of uncertainty can provide pertinent information. For example, sediment chemistry data 

that identify concentrations below the conservative 

effect level for a particular contaminant could be 

deemed safe for that contaminant. A contaminant 

concentration above the upper uncertainty level 

could be identified immediately as contaminated, 

and some degree of contamination could be 

assigned to those sediments for the individual 

contaminant. Sediments whose concentration of 

a particular contaminant falls within the degrees of 

uncertainty could require more detailed interpretation and possibly additional testing. 

6.3.2.8 Degree of Environmental Applicability 

EqP-based sediment quality criteria can be. 

applied directly to any contaminated sediment 

containing ~0.2 percent organic carbon and nonionic chemicals for which criteria are available. 

Extensive data analysis and site-specific biological 

data are not required to use sediment criteria 

developed using this method. (In some cases 

these attributes may nonetheless be desirable.) As 

a result, the EqP method can be considered environmentally applicable in some cases. Because a 

wide variety of contaminated sediment site-s exist, 

absolute statements regarding environmental 

applicability are difficult to make. However, the 

EqP method would be appropriate in many situations to predict bioavailability, bioaccumulation, 

and biological effects. 

64 
ii-E9P Approach 

6.3.2.9 Degree of Accuracy and Precision 

Bach sediment criterion value developed using 

the EqP method will have an associated degree of 

uncertainty, which will vary from criterion to 

criterion. The principal uncertainties associated 

with sediment criteria developed using the EqP 

method are those associated with partition coefficients. Hence, each developed sediment criterion 

should be and is bracketed with uncertainty, 

thereby providing decision-makers with a greater 

understanding of the meaning of the developed 

values. 

6.4 

for 

Sl-ATUS 

The method for developing sediment criteria 

nonionic organic contaminants has been 

developed and has been reviewed by the SAB on 

hvo separate occasions. Guidelines and guidance 

on the regulatory use of sediment criteria are 

under development. The method for developing 

sediment criteria for metal contaminants is being 

investigated and results are promising. The metals 

method is expected to be sufficiently well developed for peer review by 1993. 

6.4.1 Extent of Use 

Specific regulatory uses for EqP-based sediment quality criteria are being developed. A 

formal framework for the application of sediment 

criteria is not expected until EPA completes its 

effort to develop a contaminated sediment managcment strategy. The range of potentia1 applications is very large because the need for evaluating 

potentially contaminated sediments arises in many 

contexts. 

Interim sediment criteria values were developed for a variety of organic compounds. These 

values were used in a pilot study at a number of 

sites where site characterization and evaluation 

activities were conducted. The interim criteria 

were used in three ways: 

n To identify the extent of contamination 

and responsible parties; 

l To assess the risks associated with sediment contamination; and 

n To identify the environmental benefits 

associated with a variety of remedial 

options. 

A number of States have used interim and 

draft sediment criteria to evaluate the potential 

effects of several contaminants found in sediments 

in state waters. The methodologies for deriving 

sediment criteria have been used in a variety of 

situations including the evaluation of dredged 

material, Superfund site assessments, and the 

identification of appropriate cleanup levels for 

contaminated sediment sites. 

6.4.2 Extent to Which Approach Has Been 

Field-Validated 

Considerable effort has been made by EPA to 

use field sites as part of the criteria validation 

effort and to aid in designing regulatory programs. 

Table 6-1 lists ongoing and completed studies 

where SQC are being used to directly support 

sediment activities. In addition to these siles, 

there are other sites and situations (completed, 

ongoing, and planned) where the BqP is being 

applied to field situations. Although these efforts 

are not invoived with criteria development efforts, 

they do provide valuable data on the appropriateness of the EqP. 

I1 needs to be understood, however, that “field 

validation” does not describe a specific experimental protocol. The idea is to find a site that is 

contaminated with a single chemical and determine whether the benthic populations are degraded 

when the SQC is exceeded. However, there are 

practical difficulties. Such a field site contaminated with only one chemical must be found, and 

there can be no ongoing sources of the chemical 

since the exposure should be only from the se& 

ment. A gradient of chemical concentration that 

spans the SQC concentration is necessary. The 

6-7 
Sediment Classijicntion Methods Compendium 

Table 61. Ongoing rnd Completed Studies Using SM. 

sediment type must be essentially uniform in the 

gradient so that only chemical concentration is 

changing. The benthic population must be plentiful enough so that population degradation can be 

observed as the SQC is exceeded. In spite of the 

difficulties, major field efforts are presently under 

way. 

An intermediate level of field validation is 

provided by the benthic colonization experiments. 

The experimental design is desaibed above. The 

populations that develop are determined entirely 

by natural recruitment. The uniformity of sediment type is guaranteed by the experimental 

design. The experiments last from 2 to 4 months 

so that’ the sediment can properly be called a 

“natural” sediment. Three benthic colonization 

experiments have been performed using spiked 

sediments. The data analysis, which is partially 

complete, indicates that the experiments are 

consistent with the SQC for the chemicals being 

tested. 

A third type of field validation is proceeding 

as well. It is based on the notion that although it 

is not possible to prove the validity of SQC 

(continual accumulation of evidence in favor of its 

validity does not guarantee that all evidence will 

always be supportive), it is possible to prove that 

it is invalid. If sedimeuts are colleckd and the 

state of the bentbic population is evaluated relative 

to control site5 from the same region, there are 

6-8 
6-EqP Approach 

Table 6-2. SW Reid Validation Truth Table. 

four possibilities, which are arranged as a truth 

table in Table 6-2. 

The correlation of the presence or lack of 

benthic impact with exceeding or not exceeding 

the SQC is consistent but not proof of causality. 

The observation of benthic impact where the SQC 

is not exceeded can be attributed to the impact of 

other chemicals. However, if the SQC is exceeded, with a proper accounting for the uncertainty of 

SQC, and no benthic impact is observed, then the 

SQC is invalidated. The collection of these data 

is an ongoing part of the SQC development effort 

Analysis to date suggests that these data do not 

invalidate the SQC. 

6.43 Reasons for Limited Use 

The EqP method is not commonly used for 

the following reasons: 

n 

n 

n 

6.4.4 

The method was developed only recently, 

and sufficient time has not elapsed for the 

principles to be understood and used by 

others. 

Final criteria have not been issued. 

Guidance and technical support documents are in draft form and will be issued 

along with final criteria. 

Outlook for Future Use and Amount 

of Development Needed 

This method is the only procedure for derivation of sediment quality criteria that is generic 

auoss sediments, accounts for bioavailability of 

chemicals, and relates effects to specific chemicals. Therefore, EqP-based sediment quality 

criteria will be used much as water quality criteria 

are being used to define environmentally acceptable concentrations. Sediment quality criteria, 

along with sediment toxicity tests analogous to 

water quality criteria and whole-effluent toxicity 

tests, will play a major role in EPA’s management 

of contaminated sediment. 

6.5 REFERENCES 

USEPA. April 1989. Briefing report to the EPA 

Science Advisory Board on the equilibrium 

partitioning approach to generating sediment 

quality criteria. Office of Water, Regulations 

and Standards, Criteria and Standards. 

USEPA. February 1990. Report of the Sediment 

CXteria Subcommittee of the Ecological Processes and Effects Committee - Evaluation of the 

equilibrium partitioning approach for assessing 

sediment quality. A Science Advisory Board 

Report. 

USEPA. August 1991. Analytical method for 

determination of acid volatile sulfide in sediment 

(final draft). Office of Science and Technology, 

Health and Ecological aiteria Division. 

USEPA. August 1991. Technical basis for 

establishing sediment quality criteria for nonionic chemicals using equilbrium partitioning. 

Office of Science and Technology, Health and 

Ecological 0iteria Division. 

USEPA. November 1991. Proposed sediment 

quality criteria for the protection of benthic 

6-9 
St&men t Cluss$cation Methods Compendium 

organisms: Acenapththene (draft). Office of 

Science and Technology, Health and Ecological 

Criteria Division. 

USEPA. November 1991. Sediment quality 

criteria for the protection of benthic organisms: 

Dieldrin (draft). Office of Science and Technology, Health and Ecological Criteria Division. 

USEPA. November 1991. Sediment quality 

criteria for the protection of benthic organisms: 

Endrin (draft). Office of Science and Technology, Health and Ecological Criteria Division. 

USEPA. November 1991. Sediment quality 

criteria for the protection of benthic organisms: 

Pluoranthene (draft). Office of Science and 

Technology, Health and Ecological Criteria 

Division. 

USEPA. November 1991. Sediment quality 

criteria for the protection of benthic organisms: 

Phenanthrene (draft). Office of Science and 

Technology, Health and Ecological Ckiteria 

Division. 

6-10 
CHAPTER 7 

Tissue Residue Approach 

Phillip M. Cook 

U.S. Environmental Protection Agency, Environmental Research Lab-Duluth 

6201 Congdon Boulevard., Duluth, MN 55804 

(218) 720-5553, FTS 780-5553 

Anthony R. Carlson 

U.S. Environmental Protection Agency, Environmental Research Lab-Duluth 

6201 Congdon Boulevard., Duluth, MN 55804 

(218) 720-5523, FTS 780-5523 

Henry Lee II 

U.S. Environmental Protection Agency, Environmental Research Lab-Newport 

Marine Science Drive, Newport, OR 97365 

(503) 867-4042 

In the tissue residue approach, sediment 

chemical concentrations that will result in acceptable residues in exposed biotic tissues are determined. Concentrations of unacceptable tissue 

residues may be derived from toxicity tests performed during generation of chronic water quality 

criteria, from bioconcentration factors derived 

from the literature or generated by experimentation, or by comparison with human health risk 

criteria associated with consumption of contaminated aquatic organisms. The tissue residue 

approach generates numerical criteria and is most 

applicable for nonpolar organic and organometallic 

compounds. 

7.1 SPECIFIC APPLICATIONS 

7.1.1 Current Use 

Tissue residues of chemical contaminants in 

aquatic organisms, particularly fish, are frequently 

used as measures of water quality in both freshwater and marine systems. The tendency of 

organisms to bioaccumulate chemicals from water 

and food is one of the factors used in establishing 

national water quality criteria (WQC) for the 

protection of aquatic life (Stephan et al., 1985). 

Nonpolar organic chemicals, which may bioaccumulate to levels toxic to organisms or render 

organisms unfit for human food, generally will 

also be found as sediment contaminants. Hydrophobic organic chemicals preferentially distribute 

into organic carbon in sediment- and lipid in 

aquatic biota. The tissue residue approach has 

been used recently to establish the amount of 

reduction of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) concentration in Lake Ontario sediments 

necessary to attain acceptable TCDD levels in fish 

(Cook et al., 1990). The acceptable sediment 

TCDD concentration is being used as a sediment 

criterion to determine the remedial action necessary to reduce the incremental loading of TCDD 

from the Hyde Park Superfund site to Lake Ontario (Carey et al., 1989). Tissue residues of benthic 

organisms have also been used in some regulatory 

actions, such as the assessment of bioaccumulation 

potential of dredged materials (USACE, 1991). 

7.1.2 Potential use 

Although tissue residues have been used more 

commonly to determine the potential for bioaccumulation of chemical contaminants from sediments 

and dredged materials, they also provide an excellent measure of “effective exposure dose”: a measure of an organism’s actual exposure over time to 

a pollutant of concern. This exposure measure may 

be related to the dose expected at the water quality 

criterion or related directly to the potential for 

producing chronic toxic effects. Given the ability to 

measure or predict tissue residues resulting from 
Sediment Classification Methods Compendium 

exposures in contaminated sediment systems, it is 

possible to establish sediment criteria based on 

residue-toxicity effects relationships. These relationships can provide a basis for sediment criteria 

that are free of uncertainties normally associated 

with organism exposures and sediment contaminant 

bioavailability. This is especially true when in situ 

measurements provide the basis for the sediment 

residue link to the residue-toxic effect relationship. 

One example of tissue residue-toxic effects 

linkage is the relationship between the failure of 

Great Lakes lake trout (Salvelinus namaycush) to 

reproduce and bioaccumulation of TCDD and 

non-ortho substituted PCBs (Mac, 1988). Laboratory studies have shown significant mortality of 

larvae when lake trout ova contain as little as 50 ppt 

2,3,7,8-TCDD (Cook et al., 1990; Walker et of., 

1991). This residue level is found in Lake Ontario 

lake trout that have not successfully reproduced 

naturally for many years. On the basis of TCDD 

toxic equivalents for organochlorine components 

having the same mode of toxic action, residues in 

lake trout from Lakes Ontario and Michigan may 

provide a measure of the reduction in sediment 

contamination necessary to reduce fish tissue concentrations to a threshold presumed to allow reproduction. The same approach can be used for 

benthic organisms, which may have greater intersite 

variations in residue levels than do fish because of 

benthic organisms’ closer association with 

sediments. 

7.2 DESCRIPTION 

7.2.1 Description of Method 

The tissue residue approach involves the establishment of safe sediment concentrations for individual chemicals or classes of chemicals by determining the sediment chemical concentration that will 

result in acceptable tissue residues. This process 

involves two steps: (1) linking toxic effects to residues (dose-response rlationships) and (2) linking 

chemical residues in specific organisms to sediment 

chemical contamination concentrations (exposure 

relationships). Methods to derive unacceptable 

tissue residues include at least three approaches: 

• The water quality titerion-residue 

approach; 

• The experimental approach; and 

• The human health approach. 

Each of these approaches idescribed briefly below. 

Water Quality Criterion-Residue Approach-A 

rapid approach for determining acceptable concentrations of tissue residues involves establishing 

maximum permissible tissue concentrations 

(MPTCs) expected for organisms at the chronic 

water quality criterion concentration previously 

established for a specific pollutant. MPTCs, when 

not available through residue measurements obtained 

with toxicity tests used for water quality criteria, can 

be obtained by multiplying the water quality criterion by an appropriate bioconcentration factor (BCF) 

obtained from the literature. When a reliable 

empirical BCF is not available, the BCF may be 

predicted from an octanol-water partition coefficient 

or a bioconcentration kinetic model. Thus, the 

absence of a water quality criterion for a chemical 

does not eliminate this approach as long as appropriate chronic toxicity test data are available for the 

species of interest. 

Experimental Approach-Tissue residue-toxic 

effects linkages can be established through a series 

of chronic dose-responsexperiments or field 

correlations. Although this approach has the advantage of directly determining the tissue residue-toxic 

effects linkages, it can be relatively time consuming 

and costly to implement for a large number of 

pollutants. The experimental approach should be 

used to test the assumptions of the water quality 

criterion-residue approach and to supplementhe 

existing tissue residue-toxic effects database. The 

experimental work can be closely coupled with the 

experiments conducted under the bulk sediment 

toxicity test approach to deriving sediment quality 

criteria (see Chapter 3, Bulk Sediment Toxicity Test 

Approach). 

Human Health Approach-Human health risk 

from consumption of freshwater fish or seafood 

7-2 
?--Tissue Residue Apprauch 

may be used as the criterion for tissue residue 

acceptability. A sediment quality criterion for a 

specific compound can be derived by establishing 

an acceptable human risk level (e.g., an excess 

human cancer risk of 1x10”) and then back-calculating to the sediment concentration that would 

result in tissue residues associated with this level 

of risk. The human health approach can generate 

sediment quality criteria lower for carcinogenic 

compounds (e.g., PCBs, dioxins, benzo(a)pyrene) 

than those criteria derived from ecological endpoints. 

The choice of method to determine a quantitative tissue residue-sediment contamination level 

relationship depends on the specific pollutants, 

organisms, and water systems of concern, as well 

as the regulatory approach (e.g., remedial action, 

wasteload allocation, Superfund enforcement). 

The linkage between organism residue and sediment chemical concentration can be made from 

site-specific measurements of sediment-organism 

partition coefficients (Kuehl et al., 1987); fugacity 

or equilibrium partitioning model (Clark et al., 

1988); predictions of organism residues; or pharmacokinetic-bioenergetic model predictions of 

organism residues that result from uptake from 

food chain, waler, and sediment contact 

(Thornann, 1989). The residue approach works 

best for aquatic ecosystems that are at or close to 

steady state with respect to the distribution of 

chemicals between biotic and abiotic components. 

Steady-state conditions are common for most 

sediment contaminants because of their persistence 

and tendency to exert long-term rather than 

episodic bioaccumulation and toxic effects. 

7.2.1.1 Objectives and Assumptions 

The objective of this approach is to generate 

numerica sediment quality criteria based on 

acceptable levels of chemical contaminants in 

sediment-exposed biota. This objective is 

parallel to that of the water quality criteria, 

except that organism residues provide measures 

of exposure to chemical contaminants rather than 

water concentrations of contaminants. By using 

tissue residues rather than interstitial water 

concentrations to measure dose, as in the equilibrium partitioning approach (Chapter 5), this 

method does not require that the organism be at 

thermodynamic equilibrium with respect to the 

sediment contamination level. The site-specific 

residue approach is powerful because it does not 

require knowledge of bioavailability relationships for each organism in the system. All 

interaction pathways between sediment and 

organisms are incorporated in the determination 

of organism-to-sediment contamination ratios. 

These can be expressed on the basis of sediment 

organic carbon-organism lipid for hydrophobic 

organic chemicals. It is assumed that reduction 

in sediment contaminant concentrations over 

time (e.g., as a result of remedial actions, wasteload allocations) will result in parallel reduction 

in exposure, aquatic organism residues, and, 

consequently, the potential for toxic effects. It 

is further assumed that data on residue-to-toxicity relationships can be obtained from laboratory 

exposures of organisms when such data are not 

already available and that the route of exposure 

responsible for residue accumulation does not 

influence the residue-toxicity relationships. 

7.2.1.2 Level of Eflort 

Relatively little effort would be required to 

generate preliminary sediment quality criteria 

using MITCs calculated from existing water 

quality criteria and BCFs. In the absence of 

appropriate water quality criteria or BCFs, the 

level of effort depends on the availability of 

tissue residue action Ievels and the complexity 

of the sediment contaminant mitigation approach 

to be used. Relatively little effort is required to 

determine the degree to which sediment contamination concentrations must be reduced for 

single chemicals in well-mixed systems where 

fish residues are uniformly unacceptable for 

human consumption. Much more effort is 

required for systems having sediment contamination “hot spots” where resident aquatic organisms are eliminated or reduced in number due to 

a complex mixture of sediment contaminants. 

Another complexity that could increase the 

required level of effort is the presence of sediment contaminants that are readily metabolized 

7-3 
Sediment Cluss;f;cation Methods Compendium 

to chemicals of greater toxicity that are responsible for the observed adverse effects. In some 

cases, residue-toxic effects data would incorporate the effects of toxic metabolites. 

7.2.1.2.1 Type of Sampling Required 

Surface sediment samples must be analyzed 

for chemical contamina& of interest. Interstitial water composition does not need to be 

determined because the residues in biota are 

related to bulk sediment chemical composition. 

Sediment characteristics such as grain size, 

organic carbon content, and metal binding capacity are useful for defining sediment-to-biota 

relationships for different sites witbin an ecosystem. Biota sampling for residue analysis should 

include sensitive organisms when toxic effects 

are a concern or, in the absence of sensitive 

organisms, organisms whose residues will serve 

as biomarkers for establishing safe sediment 

contaminant levels. 

7.2.1.2.2 Methods 

The tissue residue approach, as discussed in 

Section 7.2, depends on determining residues in 

aquatic organisms that are unacceptable on the 

basis of toxicity to the organism or unsuitability 

for human or animal consumption as food. The 

linkage of sediment contaminant concentrations 

to organism residues is possible through a number of approaches including site-specific 

measurements, equilibrium partitioning-based 

predictions, and steady-state food chain models. 

The choice of a specific approach depends on 

the chemical of concern, the availability of 

residue-toxic effects data, the contamination 

history (in-place pollutant problem versus a 

continuing or projected sediment contamination 

problem), and the characteristics of the impacted 

ecosystem. The construction of comprehensive, 

systematic strategies for all potential sediment 

contamination assessments will be achieved 

through further research and development. 

Toxicity identification evaluation (TIE) 

procedures (see Chapter 5) complement the 

tissue-residue approach. The TIE approach is 

especialiy useful if sediment assessment begins 

without knowledge of the sediment contaminants 

that are causing toxicity or unacceptable residues 

in biota. The absence of bentbic species or 

failure of fish eggs to hatch may be attributable 

to acutely toxic, but non-residue-forming, chemicals (e.g., ammonia) in sediments. TIE procedures can distinguish between potential metal, 

nonpolar organic, polar organic, and inorganic 

chemical sources of toxicity in sediment pore 

waters or elutriates. These procedures enable a 

more complete assessment of the significance of 

residue-associated toxicity in the system. 

Once potentially toxic, bioaccumulative 

contaminants are identified, eitber in sediment or 

in aquatic organisms associated through exposure to sediments, the toxicological significance 

of site-specific sediment-to-biota contaminant 

partition factors can be assessed. Conservative 

generic sediment quality criteria can be generated from residue-toxicity relationships by assuming equilibrium partitioning between the binding 

fractions of organisms and sediments (e.g., lipid 

and sediment organic carbon for nonpolar organic chemicals). 

7.2.1.2.3 Types of Data Required 

The tissue residue method requires identification of chemicals in the sediment that are 

likely to be associated with chronic environmental effects. An indirect method for identifying 

such chemicals and their locations is to screen 

aquatic organisms for residues as in the National 

Dioxin Study (USEPA, 1987b) or the National 

Study of Chemical Residues in Fish (USEPA, 

1992), sponsored by EPA’s Office of Water 

Regulations and Standards. When toxicity data 

are not available, either laboratory dose-response 

experiments or quantitative structure-activity 

predictions can be used to establish the toxicological significance of the tissue residues. Field 

surveys that indicate the absence of sensitive 

organisms in contaminated sediment areas are 

useful for establishing sediment quality criteria, 

especially if interspecies sensitivities to the 

chemicals of concern are known. Tissue residues associated with no-effect and lowest74 
7-5s~~~ Residue Approwh 

observable-effect concentrations are needed 

when the sediment criterion is not based on a 

human health standard. 

7.2.1.2.4 Necessary Hardware and Skills 

Sediment and tissue analyses require commonly available chemical analytical capabilities. 

Some chemicals require advanced instrumental 

analytical techniques, such as high resolution 

gas chromatography/mass spectrometry. 

7.2.1.3 Adequacy of Documentation 

The use of tissue residues to establish sediment criteria on the basis of human health effects associated with ingestion of contaminated 

fish has been documented. Methods for using 

tissue residue-toxicity relationships to establish 

sediment criteria, although scientifically sound, 

have not been extensively documented. The 

various methods for predicting tissue residues in 

benthos and fish have been well documented. 

7.2.2 Applicsbillty of Method to Human 

Health, Aquatic Life, or Wildlife 

Protection 

Tissue residue measurements are directly 

applicable to human risk assessment when the 

aquatic organism is used as human food. Because of this relationship, the tissue residue 

method provides a direct link between human 

health and sediment criteria development. Tissue residues for wildlife and aquatic organisms 

can be used to assess sediment toxicity when 

there is an established exposure linkage to the 

sediment. The tissue residue approach is most 

advantageous for sediment contaminants that 

adversely impact organisms such as fish that are 

not in direct contact with the sediment or its 

interstitial water. The tissue residue approach is 

well suited to evaluating sediment quality in 

systems that have aquatic food chain connections 

from benthos to birds experiencing eggshell 

thinning or genotoxic effects. The tissue residue 

concentration serves as a quantitative measure of 

sediment contaminant bioavailability, which may 

differ as a function of ecosystem, sediment, 

water, food chain, and species characteristics. 

7.23 Ability of Method to Generate 

Numerical Criteria for Specizic 

Chemicals 

The tissue residue approach can be used to 

generate site-specific numerical criteria for nonpolar organic chemicals such as PCDDs, PCDFq 

bd PCBs. Tissue residues of aMrin/dieldrin 

(USEPA, 1980a) and endrin (USEPA, 198Ob) 

have been used to establish water quality criteria 

on the basis of human health risks. The DDT and 

PCB water quality criteria are based on toxic 

effects in birds and animals as a function of fish 

residues (USEPA, 198Oc, 198&l). Tissue residues 

of organometallic chemicals such as methyl 

mercury (USEPA, 1984) and elements such as 

selenium (USEPA, 1987a) have been used to 

establish water quality criteria and/or to predict 

toxic effects. There is some evidence to indicate 

that metal residues in sedimentdwelling aquatic 

organisms can reflect both metal bioavailability 

and potential metal toxicity. Thus, tissue residuetoxicity relationships for some elements could be 

used as an adjunct to the interstitial water equilibrium partitioning approach. 

73 USEFULNESS 

73.1 Environmental Applicability 

7.3.X.1 Suitability for Difirent Sediment Types 

There is no limitation to the suitability of this 

approach for different sediment types since the 

method is sensitive to bioavailability differences 

among sediments. When pelagic organisms are 

used to assess sediment quality, sediment variability in the water body tends to be averaged. 

7.3.1.2 Suitability for Difirent Chemicals or 

Classes of Chemicals 

This approach is most applicable to nonpolar 

organics and organometallics that bioamulate, 

are slowly metabolized, and exert chronic toxic 

7-s 
Sediment Classification Methods Compendium 

effects or present risks to human health. This 

approach also could work well for chemicals that 

are metabolized by the organism to nontoxic 

forms since the parent compound residue reflects 

this change in toxic potential. In some cases 

residues of known metabolites, which are more 

toxic than the parent compound, can be used to 

establish restdue-toxic effects relationships (Krahn 

et al., 1986). The approach is not useful for 

assessing sediment toxicity associated with nonresidue-forming toxic chemicals such as ammonia, 

hydrogen sulfide, and polyelectrolytes. Since 

there is evidence that metal residues in some 

sediment-dwelling organisms are indicative of 

both metal bioavailability and potential metal 

toxicity, sediment quality criteria for metals 

should be aided by a site-specific tissue residue 

approach. However, when biological species 

sequester metals in a nonbiologically available 

form, tissue residue-toxicity effects linkages may 

be obscured. The suitability of the method for 

evaluating additive, synergistic, or antagonistic 

effects associated with complex mixtures of 

sediment contaminants depends on the development of chemical mixture toxic dose-response 

relationships where dose is indicated by tissue 

residue levels. 

7.3.1.3 Suitability for Predicting Eflects on 

Differen& Organisms 

The tissue residue approach should not be 

limited by species unless organism residues cannot 

be obtained or toxic effects cannot be determined 

through water quality criteria or bioassays. The 

key species problem is identification of sensitive 

species for the sediment contaminants of concern. 

When adequate comparative toxicity data exist, 

residues from tolerant organisms may be used to 

infer sediment criteria for sensitive organisms that 

are not found in association with the sediment 

because of toxic effects. 

7.3.1.4 Suitability for In-Place Pollutant Control 

Evaluation of the association of site-specific 

tissue residues with sediment toxic chemical 

concentrations provides an established method for 

in-place pollutant assessment for both human 

health and ecological risks. Comparison of tissue 

residues in field-collected organisms to the MPTC 

would be a dired estimate of ecological risk. The 

use of resident or caged biota for bioaccumuiation 

potential and toxicity assessments is useful for 

detection of the most toxic sediments or monitoring of changes in toxicity following remedial 

adion. By weighing the relative toxicity of 

bioaccumulated pollutants (e.g., by using “dioxin 

equivalents”), evaluation of tissue residue concentrations can help identify the pollutants most likely 

responsible for toxicity and their additive contriiution to total sediment toxicily. This information 

could then be used to design the most appropriate 

and cost-effective mitigation response. 

7.3.1.5 Suitability for Source Control 

The tissue residue approach is well suited for 

establishing source control. Comparison of the 

existing or predicted tissue residue levels with 

MPTCs generates a quantitative estimate of the 

extent to which a given sediment exceeds or is 

below a sediment quality criterion. In conjunction 

with physical transport models, this information 

can then be used directly to determine acceptable 

discharge limits, wasteload allocations, or the 

types of remedial procedures required to achieve 

acceptable tissue residue levels. The Lake Ontario 

TCDD-Hyde Park Superfund case example described in Section 7.1.1 demonstrates the suitability of this approach for establishing source controls. The site-specific nature of this approach 

provides strong support for establishing controls 

on existing point and nonpoint sources of sediment contamination. 

7.3.X.6 Suitability for Disposal Applications 

When site-specific sediment-biota contaminant 

partition coefficients are unavailable, such as for 

evaluation of proposed disposal operations, the 

residue approach can be applied by predicting 

benthic tissue residues from steady-state toxiwkinetic bioaccumulation models or by conducting 

laboratory bioaccumulation tests on the dredged 

material. If adverse effects on fishes, wildlife, or 

7-6 
7-Tissue Residue Appraach 

human health are of concern at such disposal sites, 

it would then be necessary to apply a trophic 

transfer or equilibrium partitioning model to 

predict tissue residues in these higher trophic 

levels. When the disposal site already has sediments containing the contaminants of concern, 

residues in existing biota may be used to predict 

residue levels and toxic effects that would result 

from additional disposal of similarly contaminated 

dredged material. 

73.2 General Advantages aad Llmit~tlons 

7.3.2.1 Ease of Use 

The application of sediment quality criteria 

derived from tissue residues for assessing pelagic 

or benthic ecological effects is fairly direct. The 

measured or predicted sediment concentration 

would simply be compared to the sediment quality 

criterion derived from MFTCs. The development 

of a tissue residue toxicity database from laboratory bioassays would allow convenient am to the 

required biological effects endpoints. Chemical 

analyses of sediment, total organic carbon, and 

tissue samples for assessing existing conditions 

require routine analytical chemistry capabilities 

that do not present unique problems. One potential difficulty when using tissue residues in fieldcollected benthos to assess in-place sediments is 

the difficulty in obtaining sufficient benthic 

biomass for chemical analysis. This problem can 

be avoided by conducting laboratory bioaccumulation tests on field-collected sediment or by placing 

caged benthic organisms in the field. 

7.3.2.2 Relative Cost 

Costs associated with further development of 

the generic tissue residue approach for sediment 

quality criteria include (1) development of a 

residue-toxicity relationship database and (2) validation of the relationships between the MFTC and 

chronic impacts on aquatic organisms for different 

chemical classes of sediment contaminants. The 

cost of applying the method to a particular site, 

however, depends on the number of sediment and 

biota samples to be analyzed, the availability of 

residue-toxicity relationship data, and the difficulty in identifying sensitive organisms. The establishment of a sediment criterion based 011 fish 

residue levels acceptable for protection of human 

health generally results in low analytical costs 

when only a few reference sediment sites are 

needed to charaderize the system of concern. 

7.3.2.3 Tendency to Be Conmvalive 

This approach does not tend to be either 

conservative or it&era1 for prediction of ecological 

effecis unless the system responds in a nonlinear 

manner to reductions in sediment contaminants. 

Tn the case of nonlinearity, the tendency would 

probably be toward conservatism because of the 

greater bioavailability of more recently introduced 

sediment contaminants. When human health 

endpoints are used to generate sediment quality 

criteria, the criteria may be more strict than necessary to protect resident biota. 

7.3.2.4 Level of Acceptance 

The tissue residue approach is accepted as a 

basis for regulatory decisions such as the estabiishment of water quality criteria for the proteztion 

of aquatic life and its uses. The direct prediction 

of chronic toxic effects from measured or predicted tissue residues requires validation before it can 

be widely endorsed. Since sediment contaminants 

tend to be long-term exposure problems and can 

bioaccumulate, residues should be acceptable for 

sediment criteria development. This approach 

should be acceptable for identifying sediments 

associated with a degree of exposure which exceeds that indicated as deleterious in previous 

experiments. 

7.3.2.5 Ability to Be Implemented by 

Luhrahes with l)pical Equipmnt 

and Handling Facilities 

The tissue residue approach requires analyses 

of only sediment and tissue residues when potentially toxic sediment contaminants are known and 

residue-toxicity relationship data are available. If 

extensive laboratory work is needed to determine 

7-7 
Sediment Classification Methods Compendium 

chemical residue-chronic toxicity dose-response 

relationships for sensitive species, specialized 

aquatic toxicology capabilities are required. In 

theory, residue-toxicity-based MF+T(S can be 

obtained for all chemicals subject to water quality 

criteria development. 

7.3.2.6 Lmel of Efforr Repuired to Generate 

Results 

The level of effort depends on the number and 

nature of sediment contaminants, the compiexity 

of the contaminant distribution pattern, and the 

regulatory application of the method. Some cases 

will require relatively few analyses of tissue and 

sediment residues and no toxicity testing to apply 

the method (e.g., to remedial action decisions, 

wasteload allocations). 

7.3.2.7 Degree to which Results Lend 

Themselves lo Inlerprerahn 

Tissue residues that exceed concentrations 

considered safe for human exposure through 

seafood consumption require no interpretation 

when used to set residue-based sediment criteria. 

However, the degree of interpretation may be very 

large when evaluating ecotoxicological effects 

attributed to site-specific measurements of sediment-to-biota chemical partitioning. This interpretation problem exists for all sediment classification 

methods when applied on a site-specific basis. 

The presence of unacceptable residues in indicator 

organisms resident in or linked to an area of 

sediment contamination can be used without 

elaborate interpretation to determine compliance of 

sediments with sediment quality criteria. 

7.3.2.8 Degree of Environmen:al Applicability 

The use of site-specific tissue residues as 

quantitative exposure biomarkers eliminates 

uncertainties associated with chemical bioavailability; exposure duration, frequency, and magnitude; and toxicokinetic/bioenergetic factors. When 

the tissue residue approach is applied on a generic 

basis to generate sediment criteria for different 

chemicals, these uncertainties can be partially 

addressed through classification of sediments and 

exposure environments. 

7.3.2.9 Degree of Accuracy and Precirion 

Sediment and tissue residue chemical concentrations can be determined accurately and precisely for most chemicals. Most uncertainties in 

sediment/organism partition coefficients are due to 

biological variability. Accuracy and precision can 

be maximized through site-specific investigations 

of biological factors that influence organism 

linkage to sediment (through food chain, water, or 

direct contact) and through refinement of residuetoxicity relationships. 

7.4 STATUS 

7.4.1 Extent of Use 

Use of tissue residues to establish sediment 

criteria on the basis of human health effects has 

been documented. Tissue residues have also been 

used to derive water quality criteria for the protection of aquatic life and wildlife connected to the 

aquatic food chain. Tissue residue-toxicity data 

that may be used for d&riving numerical sediment 

quality criteria for some chemicals already exist in 

water quality criteria documents, fish consumption 

advisories, and the peer-reviewed literature. Much 

aquatic toxicology work in progress or planned for 

the future could produce the necessary data if 

residue-based dose measurements are incorporated 

into research plans. 

7.4.2 Extent to Which Approach Has Been 

Field-Validated 

Sediment TCDD contamination limits have 

been established for Lake Ontario on the basis of 

fish tissue residues. This use of tissue residue to 

generate sediment criteria has been validated 

through a steady-state model (Endicott el al., 

1989) and a laboratory bioaccumulation study 

(Cook et al., 1989) that demonstrated a linear 

relationship at steady-state between sediment 

contaminant concentration and bioaccumulated 

78 
7-Tissue Residue Approach 

TCDD in lake trout, regardless of route of uptake. 

Declines in DDT residues in fish and birds since 

its use was banned are associated with declining 

surficial sediment concentrations in the Great 

Lakes, the Southern California Bight, and elsewhere. Although other examples of studies 

validating the residue approach for single chemicals are available, its use for complex mixtures of 

chemicals in sediments to predict acceptable 

contaminant concentrations with ecosystem protection in mind has not been validated. 

7.43 Reasons for Limited Use 

Use of the tissue residue approach has been 

limited for the following reasons: 

m This method is in a developmental stage 

and has not been formally adopted by 

EPA. 

m Aquatic toxicology has only recently progressed to an understanding of residuebased dose-response relationships for sediment contaminants. 

n Regulatory agencies, including EPA, have 

not yet become committed to systematic 

establishment and application of sediment 

criteria methods. 

n The available and potentially available 

residue-based toxicity data have not been 

collated into a database for potential 

sediment criteria users. 

7.4.4 Outlook for Future Use and Amount 

of Development Yet Needed 

This method can be implemented with a 

minima1 amount of effort in many cases, especially where a single chemical or toxicologically 

related family of chemicals is of concern. Guidance documents should be written and reviewed. 

Tissue residue criteria should be accumulated 

systematically for a database. The use of this 

method in combination with other sediment 

classification methods should be considered. Field 

validation of residue-based ecologicat effects 

predictions is essential. All sediment assessment 

methods should be developed with concern for 

identification of and application to those chemicals in the aquatic environment that are iong-term 

sediment contaminants having chronic toxicity 

potential. 

7.5 REFERENCES 

Batterman, A.R., P.M. Cook, K-B. Lodge, D.B. 

Lothenbach, and B.C. Butterworth. In press. 

Methodology used for a laboratory determination of relative contributions of water, sediment and food chain routes of uptake for 

2,3,7,8-TCDD bioaccumulation by lake trout 

in Lake Ontario. Chemosphere. 

Carey, A.E., N.S. Shifrin, and A.C. Roche. 1989. 

Lake Ontario TCDD bioaccumulation study 

final report. Chapter 1: introduction, background, study description and chronology. 

Gradient Corporation, Cambridge, MA. 

17 PP. 

Clark, T., K. Clark, S. Pateson, D. Mackay, and 

R.J. Norstrom. 1988. Wildlife monitoring, 

modeling and fugacity. Environ. Sci. Technol. 22120-127. 

Cook, P.M., A.R. Batterman, B.C. Butterworth, 

K.B. Lodge, and SW. Kohlbry. 1990. Laboratory study of TCDD bioaccumulation by 

lake trout from Lake Ontario sediments, food 

chain and water. In: Lake Ontario TCDD 

Bioaccumulation Study - Final Report, Chapter 6. U.S. Environmental Protection Agency, 

Region II, New York. 

Endicott, D., W. Richardson, and D. DtToro. 

1989. Lake Ontario TCDD modeling report 

U.S. Environmental Protection Agency, hge 

Lakes Research Station, Environmental Research Laboratory Duluth, Gross Ile, MI. 

94 PPKrahn, M.M., LD. Rhodes, M.S. Myers, IX. 

Moore, W.D. MacLzod, and D.C. Malins. 

1986. Associations between metabolites of 

aromatic compounds in bile and the occurrence of hepatic lesions in English sole (fmophrys velulur) from Puget Sound, Washington. 

7-9 
Stdiment Classjficntion Methods Compendium 

Arch. Environ. Contam. Toxicol. 1561-67. 

Kuehl, D.W., P.M. Cook, A.R. Battennan, D. 

Lothenbach, and B.C. Butterworth. 1987. 

Bioavailability of polychlorinated dibenzo-pdioxins and dibenzofurans from contaminated 

Wisconsin River sediment to carp. Chemosphere 16:667-679. 

Mac, MJ. 1988. Toxic substances and survival 

of Lake Michigan salmonids: field and laboratory approaches. pp. 389401. In: Toxic 

Contaminants and Ecosystem Health. MS. 

Evans (ed). Wiley & Sons. 

Stephan, C.E., D.I. Mount, DJ. Hansen, J.H. 

Gentile, GA. Chapman, and WA. Brungs. 

1985. Guidelines for deriving numerical 

national water quality criteria for the protection of aquatic organisms and their uses. 

PB85-227040. National Technical Information Service, Springfield, VA. 

Thomann, R.V. 1989. Bioaccumulation mode1 of 

organic chemical distributions in aquatic food 

chains. Environ. Sci. Technol. 23:699-707. 

USACE. 1991. Influence of sediment potential 

of PCBs: Field studies at the Calumet Confined Disposal Facility. Environmental Effects of Dredging Notes - EEDP-O2-16, U.S. 

Army Corps of Engineers. U.S. Army Engineer Waterways Experimental Station, Vicksburg, MS. 

USEPA. 1980a. Ambient water quality criteria 

for aldrin/dieldrin. EPA 440/5-80-019. NTIS 

number PBSl-117301. U.S. Environmental 

protection Agency, Washington, DC. 

USEPA. 1980b. Ambient water quality criteria 

for endrin. EPA 440/S-80-047. NTIS number 

PB81-117582. U.S. Environmental Rote&on 

Agency, Washington, DC. 

USEPA. 198Oc. Ambient water quality criteria 

for DDT. EPA 440/S-80-038. NTIS number 

PB81-117491. U.S. Environmental Protection 

Agency, Washington, DC. 

USEPA. 1980d. Ambient water quality criteria 

for polychlorinated biphenyls. EPA 440/S-80- 

068. NTIS number PB81-117798. U.S. 

Environmental Protection Agency, Washington, DC. 

USEPA. 1984. Ambient water quality criteria for 

mercury. EPA 440/S-84-026. NTIS number 

PB85-227452. U.S. Environmental Protection 

Agency, Washington, DC. 

USEPA. 1987a. Ambient water quality criteria 

for selenium. EPA 440/5-87-006. NTIS 

number PB88-142237. U.S. Environmental 

Protection Agency, Washington, DC. 

USEPA. 1987b. The national dioxin study. 

Tiers 3,5,6, and 7. EPA 440/4-87M)3. U.S. 

Environmental Protection Agency, Office of 

Water Regulations and Standards, Washington, DC. 

USEPA. 1992. National study of chemical 

residues in fiih. 2 vols. EPA 823-R-92- 

008a,b. U.S. Environmental Protection Agency, Office of Science and Technology, Standards and Applied Science Division, Washington, DC. 

Walker, MK, J.S. Spitbergen, J.R. Olson, and 

R.E. Perterson. 199 1. 2,3,7,8-Tetrachlorodibenzo-p-dioxin toxicity during early life 

stage development of lake trout (Solwlinur 

namaycush). Can. I. Fish. Aqua. Sci. 48:875. 

7-10 
CHAPTER 8 

Freshwater Benthic Macroinvertebrate 

Community Structure and Function 

Wayne S. Davis 

U.S. Environmental Protection Agency Region V, Environmental Sciences Division 

77 West Jackson (SQ-14J), Chicago, IL 60604 

312/FTS 886-6233 

Joyce E. Lathrop 

College of DuPage, Division of Natural Sciences 

22nd at Lambert Road, Glen Ellyn, IL 60137 

The community, or assemblage, structure and 

function of benthic macroinvertebrates is used extensively to evaluate the quality of water resources and 

characterize causes and sources of impacts in lotic 

(flowing water) and lentic (standing water) freshwater 

ecosystems. (Marine benthic community structure is 

discussed in (Chapter 9.) Benthic macroinvertebrates 

are relatively sedentary organisms that inhabit or 

depend on the sedimentary environment for their 

various life functions. Therefore, they are sensitive to 

both long-term and short-term changes in habitat, 

sediment, and water quality. This chapter discusses 

assessment of benthic macroinvertebrates to determine 

sediment quality in conjunction with an integrated 

approach for assessing the quality of the water 

resources. This integrated approach uses sediment 

chemistry, sediment toxicity, habitat quality, and benthic macroinvertebrate community (assemblage) 

structure and function to evaluate sediment quality, 

similar to the approaches now used to evaluate 

surface water quality. The structural assessment 

relates to the numeric taxonomic distribution of the 

community, and the functional assessment involves 

trophic level (feeding group) and morphological 

assessment. This chapter addresses the specific 

benthic community assessment methods that are 

available, or being developed, to complementhe 

chemical and toxicological portions of the sediment 

quality assessment. 

8.1 SPECIFIC APPLICATIONS 

8.1.1 Current Use 

Freshwater benthic macroinvertebrate communities are used in the following ways to assess the 

quality of the water resource (sediments, water, and 

habitat): 

• 

• 

• 

• 

• 

8.1.1.2 

Identification of the quality of ambient 

sites through a knowledge of the pollution 

tolerances and life history requirements of 

benthic macroinvertebrates; 

Establishment of criteria and standards 

based on community performance at 

multiple reference sites throughout an 

ecoregion or other regionalization categories; 

Comparison of the quality of reference (or 

least impacted) sites with test (ambient) 

sites; 

Comparison of the quality of ambient 

sites with historical data to identify temporal trends; and 

Determination of spatial gradients of contamination for source characterization. 

Ecological Uses 

Benthic macroinvertebrate community (assemblage) structure and function assessments have 

many different applications. Site-specific knowledge of surface water quality, habitat quality, 

sediment chemistry, and sediment toxicity provide 

the best context in which to interpret benthic 
Sediment Classification Methods Compendium 

community assessment data. The objectives of 

each particular study should determine the types of 

related data necessary. Alone, benthic macroinvertebrates can be used to screen for potential 

sediment contamination based on spatial gradients 

in community structure, but they should not be 

used alone to definitively determine sediment 

quality. Benthic macroinvertebrate data must be 

integrated with other available data to determine 

sediment quality. Benthic macroinvertebrate often 

provide the most important piece of information on 

sediment quality. Care must be exercised to 

collect representative samples to minimize problems with data interpretation due to natural variations. For example, collections should not be 

made after floods or other physical disturbances 

that may physically alter or remove benthic assemblages. 

Benthic macroinvertebrate community structure 

and function have been used extensively to characterize freshwater ambient conditions and impacts 

from various sources. Documented changes in 

benthic community structure have resulted from 

crude oil exposure in ponds and streams (Rosenberg and Wiens, 1976; Mozley, 1978; Mozley and 

Butler, 1978; Cushman, 1984; Cushman and Goyert, 1984) and heavy metal contamination of lake 

sediments and streams (Winner et al., 1975,1980, 

Wentsel et al., 1977; Moore et al., 1979; Wiederholm, 1984a, 1984b; Waterhouse and Farrell, 

1985). Benthic macroinvertebrates have been used 

extensively to identify organic enrichment in lentic 

systems (Cook and Johnson, 1974: Krieger, 1984; 

Rosas et al., 1985) and lotic systems (Richardson, 

1928; Gaufin and Tarzwell, 1952; Hynes, 1970; 

Hilsenhoff, 1977, 1982, 1987, 1988). Benthic 

community responses to pesticides (van Dyk et al., 

1975; Webb, 1980; Penrose and Lenat, 1982; 

Yasuno et al., 1985), acid- and mine-stressed lotic 

environments (Simpson, 1983; Armitage and 

Blackburn, 1985), thermally stressed water bodies 

(Crossman et al., 1984), and urban and highway 

runoff impacts (Smith and Kaster, 1983; Dupuis et 

al., 1985; Denbow and Davis, 1986) have also 

been documented. Chironomidae (midge) larvae 

were even found to transport substantial amounts 

of PCBs from contaminated sediments to the 

terrestrial environment (Larsson, 1984). 

8.1.1.2 Regulatory Uses 

Assessment of benthic macroinvertebrate community (assemblage) structure and/or function has 

been used as a regulatory tool for a number of 

years (Davis, 1990). In 1987, USEPA hosted the 

First National Workshop on Biological Monitoring 

and Criteria (USEPA, 1988a, 1988b), which addressed the use of benthic macroinvertebrates, as 

well as fish, in EPA and State regulatory programs. This workshop formally initiated EPA’s 

efforts toward development and implementation of 

“biological criteria” based on benthic macroinvertebrate, fish, and habitat assessments. These 

biological criteria, which have been predominantly 

based on the macroinvertebrates, are designed to 

determine whether a specific water body or water 

body segment is meeting its designated use for 

aquatic life (i.e., water quality standards). 

EPA requires the development of biological 

criteria and adoption by States into their water 

quality standards by September 30,1993 (USEPA, 

1991a, 1990b). This requirement has been supported by a formal policy (USEPA, 1990c), program guidance (USEPA, 1992a), and technical 

guidance and support documents (USEPA, 1991a, 

1991b, 1991c, 1991d, 1991e, 1992b, 1992c). 

Several States currently use benthic macroinvertebrates as a regulatory tool, either alone or in 

combination with other ecological parameters 

(Ohio EPA, 1990, USEPA, 1991c, 1991e). 

USEPA also supports the use of benthic macroinvertebrates as a primary environmental indicator 

for surface. waters that EPA should use to track 

compliance with Clean Water Act objectives (Abe 

et al., 1992; USEPA, 1990d, 1990e). 

Under the Clean Water Act, as amended in 

1987, benthic macroinvertebrates are used for the 

following: 

• Measurement of the restoration and maintenance of biological integrity in surface 

waters (section 101); 

• Development of water quality criteria based 

on biological assessment methods when numerical criteria for toxicity have not been 

established [section 303(c)(2)(B)]; 

8-2 
PFreshwatcr Benthic Macroinvert&afe Community Sftudure and Fun&h 

m Production of guidance and criteria based 

on biological monitoring and assessment 

methods [section 304(a)(8)]; 

n Development of improved measures of the 

effects of pollutants on biological integrity 

(section 105); 

n Production of guidelines for evaluating 

nonpoint sources (NPS) [seciion 304(f)]; 

n Listing of waters that cannot attain designated uses without additional NPS 

controls (section 319); 

n Listing of waters unable to support balanced aquatic communities [section 

304(l)]; 

m Assessment of lake trophic states and 

trends (section 314); 

n Production of biennial reports on the 

extent to which waters support balanced 

aquatic communities [section 305(b)]; 

ad, 

n Determination of the effect of dredge and 

fill disposal on balanced wetland 

communities (section 404). 

Benthic macroinvertebrates and biological 

criteria have also been used to evaluate on-site 

and off-site ecological impacts from hazardous 

waste sites. Environmental assessment of a 

Superfund site is done in accordance with EPA’s 

responsibility to protect public health and the 

environment under the Comprehensive Environmental Response, Compensation, and Liability Ad 

of 1980 (CERCM) as amended by the Superfund 

Amendments and Reauthorization Act of 1986 

(SARA). The regulation that enables EPA to carry 

out its responsibilities under CERCLA/SA.&% is 

the National Contingency Plan (NCP). 

The NCP calls for the identification and 

mitigation of environmental impacts of these sites 

and the selection of remedial actions that are 

“protective of environmental organisms and 

ecosystems.” Federal and state laws and rcgdations that aid in this process are potentially “applicable or relevant and appropriate requirements” 

(ARARs). Compliance with these laws and 

regulations increasingly requires that the site’s 

ecological effeds be evaluated and measures be 

taken to mitigate those adverse cffeds. 

The Clean Water Ad, as amended by the 

1987 Water Quality Ad, is another ARAR and 

major federal regulation that requires the maintenance and restoration of the chemical, physical, 

and biological integrity of the Nation’s waters. 

Most Superfund sites potentially affect surface 

waters and need to be assessed for both on-site 

and off-site effects. A detailed discussion of the 

legal and technical requirements for environmental 

assessments at Superfund sites can be found in 

EPA’s Risk Assessment Guidance for Super 

Environmental Evaluation Manual (USEPA, 

1989a). As EPA focuses on watershed and water 

body impacts regardless of the programmatic 

sources and causes, the use of benthic macroinvertebrates for assessing the health of surface 

water systems will increasingly become important. 

8.1.2 Potential Use 

The use of benthic macroinvertebrates to assess 

sediment contamination will be most successful 

when combined with sediment chemistry and 

toxicity results, as in the “integrated” Sediment 

Quality Triad approach (see Chapter 10). Benthic 

macroinvertebrates will best indicate in-place 

pollutant control needs through a site-specific 

knowledge of surface water quality, habitat quality, 

and sediment chemistry and toxicity. Habitat 

quality assessments will help establish reasonable 

expectations for benthic community structure and 

function. Alone, benthic macroinvertebrates can be 

used to screen for potehtial sediment contamination 

and source identification by displaying spatial 

gradients in community structure, but they should 

not be used alane to deftitively determine sediment quality or to develop chemical-specific guidelines. Benthic macroinvertebrate data must be 

integrated with other available data to determine 

sediment quality as well as the quality of the 

overall water resource. 

8-3 
Sediment Classfjcratitm Methods Compendium 

8.2 DESCRIPTION 

8.2.1 Description of Method 

The benthic macroinvertebrate community 

structure and function assessment involves the 

following steps: 

(1) 

(2) 

(3) 

(4 

5) 

(6) 

(7) 

Establishment of data quality objectives, 

selection of sample sites and frequency of 

collection in Quality Assurance Program 

Plan; 

Collection of benthic macroinvertebrates in 

the field (artificial or natural substrates); 

Sorting the organisms from debris (field or 

laboratory); 

Identification to the lowest taxon necessary 

(varies depending on the study objectives); 

Multimetric or composite index quantification (e.g., taxa richness, number of individuals, indicator organism count, structural 

indexes and ratios, functional characteristics of taxa); 

Assessmenl of the relationship with other 

environmental measurements including 

numeric habitat quality assessment (e.g., 

correlations, habitat requirements) and 

expectations; 

Comparison with a local or regional “reference” site (e.g., similarity indexes, nonparametric analyses); and 

(8) Complete documentation of the study 

methods, results, database management, 

and discussion of the relevance of the data. 

8.2.1.1 Objectives and Assumptions 

lbe primary objective of benthic macroinvertebrate community (assemblage) structure and function analyses is to provide data and information to 

assist in determining the quality of the sediment/water environment. This determination can 

then be used for the purposes descriied above in 

Section 8.1 (Specific Applications). 

It is assumed that benthic macroinvertebrates 

can provide consistent and accurate assessments of 

sedimenlkater quality at a given sample location or 

water body. Specifically, the following assump 

tions are implicit in this objective: 

m 

n 

m 

m 

n 

The benthic macroinvertebrates are relatively sedentary, especially compared to 

fEh communities, and they depend on the 

sedimentary (or benthic) environment for 

their life functions. 

Chemical and physical perturbations of the 

sediments or bottom waters affect benthic 

macroinvertebrates since they are dependent on the benthic environment for completion of their life cycles, and they are 

therefore sensitive to changes in sediment 

and water quality. 

Bcnthic macroinvertebrates physically 

interact with the sediments to cause c&emical exchange between the sediment and 

the overlying water, and therefore tend to 

reflect sediment quality as well as water 

quality. 

Minimum habitat quality exists below 

which the community structure and function will perform poorly regardless of the 

chemical contaminants present or not 

present. 

The optimal use of benthic macroinvertebrates as sediment quality indicators is as 

part of an integrated sediment quality assessment approach using sediment 

chemistry, sediment toxicity, and benthic 

community structure and function. 

Equally important assumptions apply to actual 

benthic macroinvertebrate sampling strategy, a&ction, identification, data reduction, interpretation of 

results, and report preparation. It is assumed that 

all U.S. EPA-supported studies have an adequate 

8-4 
8-Freshwater Benthic Macroi.nvert&rate Community Structure and Function 

Quality Assurance Project Plan (QAPP) and that all 

benthic macroinvertebrate community data are 

reproducrble and collected in a manner to minimize 

data interpretation problems with natural variations; 

the methods Ipust be consistent within each study. 

Specific QA procedures that should be established 

early in benthic macroinvertebrate community 

studies include the following: 

Rationale for sample location selection; 

Sample collection methods, sorting, and 

storage procedures; 

Taxonomic proficiency evaluations using 

either U.S. EPA check-samples from Cincinnati-ERL or state-developed checksamples, in addition to voucher collections from each study ares and a list of 

the taxonomic references used, 

Multimetric data analysis techniques used 

to objectively assess the data, including 

the structural and functional measures; 

and 

Nonparametric or parametric (as appropriate) statistical methods used to compare 

site results. 

Each Regional U.S. EPA Quality Assurance 

Office can provide the details of QAPP requirements. Further discussion of quality assurance 

measures can be found in Klemm et al. (1990), 

Bode (1988) Ohio EPA (1989b), and Stribling 

(1991). 

8.2.1.2 Level of Eflort 

The level of effort required lo conduct freshwater benthic macroinvertebrate community 

studies is comparable with chemical/physical 

water quality measurements and bioassays and has 

been thoroughly discussed in Plafkin et al. (1989) 

and Ohio EPA (1990a). However, rapid benthic 

community assessment techniques can range from 

1 to 5 hours per site if laboratory identifications 

are not required (Plafkin et al., 1989). As expected, the greatest time expenditure is in the travel to 

and from the site and in the sorting and identification of the organisms. 

Separating the organisms from debris and 

sorting the organisms into taxonomic categories 

can take up to 15 hours per sample, with an 

additional 12 hours for identification, for very 

enriched sites with high numbers of individuals 

among several taxa. In such extreme situations, 

subsampling may be preferred. More typically, 

the time spent would be about 3 hours for sorting 

(more time for dredge and artificial substrate 

samples and less time for dip-net samples), 

2 hours for preparing the samples (e.g., clearing 

and then mounting the chironomids on microscope 

slides), and 6 hours for identifying the organisms 

to the lowest possible taxonomic level. An experienced taxonomist with appropriate keys may 

average only 2-4 hours per site. This typical time 

equates to about 11 hours per site after the samples have been collected. These estimates are 

only a general guide to the time it may take to 

perform the identifications and are meant to help 

assess potential or actual project costs. 

8.2.1.2.1 Type of Sampling Required 

The specific sampling methods to be used are 

dictated by the study needs. Debate will continue 

regarding the use of “quantitative” and “qualitative” sampling methods, but each method is 

acceptable contingent upon bow well it will satisfy 

study objectives, reproducibility of the data, and 

consistency of collection. Typically, benthic 

macroinvertebrate data are quantified by the 

surface area of the sampler or sediment being 

collected. However, benthic macroinvertebrates 

can be quantified in other ways depending on the 

objectives of the study. For example, if the 

objective is to determine the number and types of 

taxa in a study area, rather than the number of 

individuals within each taxon, then using a dip-net 

in various habitats within the study area until no 

new taxa are encountered could be considered 

quantitative with relation to the number of taxa 

and time expended. Examples of programs using 

data quantified by methods other than surface area 

of the sampler or substrate include those described 

8-5 
S&nent Cluss~icntM1 Methods Compendium 

by Pollard (1981), Hilsenhoff (1982, 1987,1988), 

Cummins and Wilzbach (1985), Bode and Novak 

(1988), Cummins (1988), Hite (1988), Lenat 

(1988), Maret (19881, Penrose and Qverton 

(1988), Plafkin et al. (1989), and Sbakelford 

(1988). The success of each sampling effort 

depends on a thorough understanding of the data 

quality objectives of that study and the implementation of a quality assurance program. 

8.2.1.2.2 Methods 

EPA (Klemm et al., 1990) recently published 

Macroinvertebrate Field and Laboratory Methocis 

for Evaluating the Biological Integrity of Swface 

Waters, which thoroughly addresses methodology. 

Most state environmental regulatory programs 

have a Quality Assurance Project Plan describing 

the field methods and standard operating procedures for collecting and evaluating benthic maaoinvertebrates (Bode, 1988; Illinois EPA, 1987; 

Ohio EPA, 1989a, 1989b). This information 

should be obtained to ensure acceptance and 

comparability of study results with those obtained 

by the state agency. If this information is not 

available, then field methods and standard operating procedures from other existing programs 

should be used. Since several different collection 

and analysis methods are used throughout the 

country depending on water body type (lotic vs. 

lentic), habitat type, substrate type, and familiarity 

witb specific methods, it is not practical to recommend any single sampling method. The general 

quality assurance requirements the use of any one 

particular method is that the method produce data 

that are reproducible, consistently used within the 

program, and applicable by different investigators 

(Klemm et al., 1990). 

Methods Summary-h soft freshwater sediments 

the most common method used to collect benthos 

is with a grab sampler such as a Ponar (15 x 15 

cm or 23 x 23 cm) or Ekman dredge (15 x 15 cm, 

23 x 23 .an, or 30 x 30 cm), each of which 

provides a quantitative sample based on the 

surface area of the sampler. The smaller of the 

surface area sizes are most commonly used for 

freshwater studies because of their relative ease of 

manipulation. Ihe Ekman dredge is not as effective in areas of vegetative debris, but is much 

lighter than the Ponar and easier to use in softer 

substrates. Artificial substrates (Hester-Bendy 

using several 3-inch plates and spacers attached by 

an eyebolt; or substrate/rock-filled baskets) provide a consistent habitat for the benthos to colonize in both soft-bottomed and stony areas. 

Artificial substrates can be used in almost any 

water body and have been successfully used to 

standardize results despite habital differences 

(APHA et al., 1989; DePauw, 1986; Hester and 

Dendy, 1962; Ohio EPA, 1989b; Rosenberg and 

Resh, 1982, 1991), but the major drawback to 

using the artificial substrates is the 4- to 8-week 

period for instream colonization. This would 

require at least two visits for each study site-one 

to place the samplers and one to remove them. 

A variety of methods for sampling hardbottomed lotic systems are available. Colonization 

of substrates and comparisons of the artiticial and 

natural substrate methods have been described 

(Beckett and Miller, 1982; Chadwick and Canton, 

1983; Crossman and Cairns, 1974; Lenat, 1988; 

Ohio EPA, 1989b; Peckarsky, 1986; Plafkin et al., 

1989; Shepard, 1982). If quantification by sediment or sampler surface area is needed, a Surbertype square-foot sampler (Surber, 1937, 1970) 

with a #30-mesh (0589-mm openings) can be 

used. The traveling kick-net (or dip-net) method, 

also using a #30-mesh net, can be used to quantify 

the sample collected by the amount of time spent 

sampling and the approximate surface ares sampled (Pollard, 1981; Pollard and Kinney, 1979). 

The Surber-type and kick-net methods can each be 

used to provide consistent, reproducible samples, 

but both are limited to wadable streams. The 

Surber sampler’s optimal effectiveness is limited 

to riffIes, whereas kick- or dip-net sampling can 

be used in all available habitats. Although dip-net 

samplers have been effectively used to sample 

riffles and other relatively shallow habitats to 

determine taxa richness, presence of indicator 

organisms, relative abundances, similarity between 

sites, and other information, they do not provide 

definitive estimates of the number of individuals 

or biomass per surface area. 

8-6 
8-Fmshwater Benthic Macroin&&rate Community Sfructure and Funcfion 

For sediment evaluations of lotic systems, a 

combination of artificial substrate (e.g., HesterDendy) and natural substrate (dip-net) sampling is 

recommended. ‘TEis combination allows comparison of the benthos communities independent of 

habitat so that sediment/water quality effects can 

be better assessed. 

Slmpling Strategy--Sampling strategies have 

been addressed by KIemm et al. (1990), Millard 

and Lettenmaier (1986), Plafkin et al. (1989), 

Rosenberg and Resh (1991), Sheldon (1984), and 

USEPA 199Ob, 199Oc). Special monitoring strategies have been prepared for EPA’s Environmental 

Monitoring and Assessment Program (EMAP), 

which employs a probabilistic sample design 

(USEPA, 1991f); the intensive watershed surveys 

of the U.S. Geological Survey (Leahy et al., 

1990); and forestry activities in the Pacific Northwest (USEPA, 1991g). Regardless of the study 

objectives for regulatory use, reference (leastimpacted) sites will be required for comparison 

with the results from test (ambient) sites. Reference sites can be established on a site-specific or 

regional basis. It is preferable to use a regionalization approach because the level of confidence 

in the results is greater using an increased number 

of reference sites, which allows for a verification 

that the sites truly are least-impacted reference 

sites. Regionalization (ecoregions, watersheds) 

has been successfully used in a number of State 

programs to support biological criteria deveiopment for benthic macroinvertebrates (Gallant et 

al., 1989, Ohio EPA, 1990, Arkansas DPCE, 

1987, Hughs et al., 1990, USEPA, 1991c, 1991e). 

When using site-specific reference sites to 

detect spatial differences in sediment/water quality, or to characterize sources of pollution, the best 

strategy is to collect samples in similar habitats 

upstream and downstream of suspec!ed pollution 

sources or other areas of interest for ambient 

monitoring such as high-quality or wild and scenic 

streams (USEPA, 1992b). A minimum of two 

upstream sites and three downstream sites of the 

suspected pollutant source(s) should be sampled, 

however, many programs are limited to only one 

upstream site and one or two downstream sites. If 

habitats vary too widely, then artificial substrates 

should be placed at each site, with multihabitat 

dip-net sampling done when the substrates are 

placed instream and retrieved, to complement the 

artificial substrate data. 

To best detect temporal trends, a fued station 

network should be established near the area of 

interest and sampled consistently at least one 

season each year. A reference location should 

also be sampled at the same times to ensure that 

differences found in the results can be attributed 

to changes in water quality near the site. It is 

strongly recommended that a set of reference sites 

be developed withirl each ecoregion (or by other 

regional&&ion methods) and that those reference 

sites be sampled seasonally to better understand 

site-specific seasonal variability. Sampling should 

be done each year during similar flow conditions 

and should not be conducted for at least 1 or 2 

weeks after a major rainfall because of the potential for physical disturbances of the substrate 

resulting in potentially lower biological integrity 

ratings. 

Seasonal distributions are always a conceru 

for ensuring the collection of a representative 

sample. Therefore, routine sampling or monitoring is optimal during the seasons indicated in 

Platlcin et al. (1988), and long-term monitoring 

should strive for consistent sampling seasons. The 

benthic macroinvertebrate discussion group at the 

1987 National Workshop on Instream Biological 

Monitoring and Criteria agreed that the biologically optimal time of year for sampling in iotic 

systems was during the latter part of the season(s) 

that demonstrate a stable base-flow (normal flow) 

and temperature regime (Davis and Simon 1988). 

Sample R~plicatio~ample replication is a 

component of a good Quality Assurance Program 

Plan. Recommendations and discussion regarding 

sample replication can be found in Plafkin et al. 

(1989), Klemm et al. (1990), and USEPA (1992b). 

Statistical derivation of the number of samples 

required to decrease the variability of the data 

have been discussed by Green (19781, Merritt et 

al. (1984), Resh and Price (1984), and Klemm et 

al. (1990). These methods generally rely on a 

prior knowledge of the variability of the data. 

This prior knowledge is often not available nor 

8-7 
Sediment CZass$cation Methods Compendium 

practical to obtain from a programmatic view 

(e.g., the cost of initial sampling to estimate 

variability and required number of replicates may 

be prohibitive). Another problem with statistically 

determining the number of samples needed is the 

assumption that the data follow a specific distribution such as normal or lognormal, which is not 

necessarily true for biological samples. Also, the 

variability, as measured by the variance or 

standard deviation, could be different for each 

descriptive index analyzed (number of taxa versus 

number of individuals, etc.). 

Field Methods-Field sampling methods have 

been adequately addressed by many manuals, 

including the new USEPA macroinvertebrate field 

and laboratory manual (Klemm et al., 1990), the 

ASTM methods for sampling benthos (ASTM, 

1988), Ohio EPA’s Field Methods Manual (Ohio 

EPA, 1989b), Standard Methods (APHA et al., 

1989), USEPA’s Rapid Bioassessment Protocols 

(Plafkin et al., 1989), and USEPA’s Superfund 

Field Compendium (USEPA, 1987). The following decisions will need to be made once the 

sample gear is chosen: 

m Whether samples will be picked from 

debris and sorted in the field, 

n Which preservative should be used; 

n Whether a stain (rose bengal) will be 

added to the sample to facilitate separating the organisms from debris; 

l Whether the samples need to be shipped 

and whether they require a chain-of-custody form (as in Superfund samples); and 

8 The type of sample containers and labeling of tbe containers required. 

Sorting--There are many discussions elsewhere 

of techniques for sample sorting and preparation 

of slides for identification. Klemm et al. (1990), 

Merritt el al. (1984), Pennack (1978) and APHA 

ef al. (1989) offer excellent guidance for sample 

sorting. Hynes (1970, 1971) stated that the 

earlier stages of benthos are retained by a 0.2-mm 

mesh size (approximately the size of a W75 standard sieve), and APHA et of. (1989) and Klemm 

et 01. (1990) defined the benthos by a mesh size 

of 0.595 (standard sieve #30), which is now 

standard practice. However, some types of chironomidae and other small benthos pass through the 

#30-mesh sieve but are be retained by the #40- 

mesh sieve. It is therefore recommended that 

samples be passed through a #30-mesh sieve and 

that the materials washed through be passed 

through a &to-mesh sieve; the material retained in 

both sieves should ihen be sorted (Ohio EPA, 

1989b). Once the material is washed through the 

sieves the organisms should be separated from the 

vegetation and other debris in a white enamel pan. 

As the materials are separated, the organisms can 

be placed in different vials for the major taxa. 

Taxonomy--The level to which the taxonomy 

should be taken is dependent on the objectives of 

the study. For a system reconnaissance or screening survey, it is generally not necessary to go 

beyond the family level (Hilsenhoff, 1988; Illinois 

EPA, 1987; Plafkin et of., 198% Resh, 1988). For 

studies attempting to identify designated use 

impairment and/or evaluate impacts from a specific source., the recommended minimum level of 

taxonomic detail should follow the list by Ohio 

EPA (1989b). Ohio EPA has successfully implemented numeric biocriteria based on this taxonomic detailing. ‘Ihis strategy is to expend the effort 

to differentiate those taxa which are better water 

resource quality indicators and for which taxonomic keys and expertise are readily available. 

The level of taxonomic detaifing must be consistent within the program and applied for each study 

site. Species-level identifications for all organisms 

are not necessary for a successful program, and 

they commonly depend on the availability of local 

keys. General keys available for genus-level 

identifications include Merritt and Cummins 

(1984) for insects, Peckarsky et al. (1990) for 

insects and other invertebrates, Pennack (1978, 

1989) for all common invertebrates, Wiederholm 

(1983) for midges, and Klemm (1985) for annelids 

(oligochaetes and leeches). Klemm et al. (1990) 

provide an excellent list of taxonomic references 
&-Freshwater Benfhic Macroinvertebrate Community Strudure and Function 

for other general and specific uses. Regional U.S. 

EPA or state biologists should be contacted to 

determine which of the hundreds of other taxonomic keys are available for specific taxa, both 

nationally and regionally. 

8.2.1.2.3 Types of Data Required 

The types of data analyses that are required to 

meet program objectives directly affect the types 

of data required. A list of the families of taxa 

present may be sufficient to meet some program 

objectives. Under other circumstances, specieslevel taxonomy and enumerations may be required. The necessary data required to conduct 

different types of analyses can be obtained from 

the following discussion of data analysis methods. 

One of the most inconsistent and perplexing 

aspects of a freshwater benthic macroinvertebrate 

community assessment is the numeric representation and analysis of the data collected. Structural 

community measures such as richness values, 

diversity and biotic indexes, and enumerations 

have been used almost exclusively. Indicator 

organisms have been used to establish many of the 

biotic indexes but also have the potential to 

differentiate among types of impacts. Recently, 

functional community measures based on feeding 

groups such as shredder, collector, scraper, and 

predator (Cummins and Merritt, 1984) have 

gained wider application and acceptance due to 

their sensitivity in detecting system perturbation 

on food resources. Sediment and water quality 

assessments based on the benthic macroinvertebrate community should use a complementary mix 

of both structural and functional measures. It is 

strongly recommended that a multimetric lechnique be used (Plafkin et al., 1989; Ohio EPA, 

1990a) so any single index value or observation 

will not substantially influence the results. Discussions of various data analysis techniques can 

be found in Hawkes (1979), Cairns (1981), 

Klemm et al. (1990), Washington (1984), and 

Resh and Jackson (1990). 

Composite Iodexes&mposite or multimetric 

indexes combine selected structural or functional 

measures, or “metrics,” in a cumulative scoring 

system, as was done with the Index of Biotic 

Integrity (IBI) for the fish community (Karr et al., 

1986). These composite, or multimetric, indexes 

are highly recommended and are among the most 

used assessment techniques for development of 

biological criteria for both benthic macroinvertebrates and fish. 

Karr and Kerans (1992) provide an outstanding discussion of the process of developing metrics proposed for use in an invertebrate IBI. They 

evaluated 28 potential metrics for inclusion and 

have eliminated 10 from further consideration. 

The metrics fail into three categories: taxa richness and community composition, trophic and 

functional feeding group, and abundance. 

Ohio EPA (1989b, 1990a) successfully deveioped a similar index for invertebrates using the 

following 10 structural metrics, adjusted for 

drainage area size with each ecoregion, to derive 

a final Invertebrate Community Index (ICI) score: 

(1) Total number of taxa; 

(2) Total number of mayfly taxa; 

(3) Total number of caddisfly taxa; 

(4) Total number of dipteran taxa; 

(5) Percent mayflies; 

(6) Percent caddisflies; 

(7) Percent Tribe Tanytarsini midges; 

(8) Percent other dipterans and non-insects; 

(9) Percent tolerant organisms; and 

(10) Total number of qualitative EPT taxa. 

The ICI score is part of Ohio EPA’s numeric 

biocriteria for designated use attainment, and it 

was developed using artificial and natural substrate data for 232 “least-impacted” reference sites. 

A statistical validation of the ICI using a factor 

analysis technique showed high correlations 

between the factor analysis scores and the ICI 

8-9 
Se&rent ClassijiiXltion Methods Compendium 

scores and little redundancy between the metrics 

(Davis and Lubin, 1989). 

U.S. EPA (Plafkin er al., 1989) developed a 

composite in&x for rapid assessments in lotic 

systems using the following two functional and six 

structural metrics: 

(1) Taxa richness; 

(2) Modified Hilsenhoff biotic index; 

(3) Ratio of scrapers and filtering collectors 

(functional); 

(4) Ratio of EPT and Chironomidae abundances; 

(5) Percent contribution of dominant taxon; 

(6) EPT index; 

(7) Community similarity index; and 

(8) Ratio of shredders to total number of 

organisms (functional). 

These Rapid Bioassessment Protocols (RBPs) 

recommend conducting single-habitat (riffle) dipnet sampling. The scores are based on a percentage of the metric values found at a reference site, 

rather than comparison of the results based on 

“optimal” values for each metric. Modifications to 

the RBPs can include use of multiple reference 

sites. The RBPs are flexible and can be modified 

for different geographical locations, as evidenced 

by the use of different metrics in Arkansas (Shakelford, 1988) and New York (Bode and Novak, 

1988). The success of the RBPs is in the use of 

the composite index for rapid assessments that 

allows for three levels of taxonomic work (i.e., 

order, family, or genus/species levels). Order and 

family taxonomy do not require laboratory taxonomy and may be done in the field. The RBPs 

normally use single-habitat (riffle) sampling and a 

lOO-organism count in the field. However, they 

can be adapted for most program uses; for example, by employing multihabitat sampling and/or 

various count limitations. To be applicable to a 

state’s program, the RBPs should undergo a 

rigorous validation effort within that state. 

Diver&y Indexes-When diversity indexes were 

introduced, they were used widely because of their 

ability to reduce the complex benthic community 

measurements into a single value that could be 

used by nonbiologist decision-makers. Diversity 

indexes are based on measuring the distribution of 

the number of individuals among the different 

taxa, and use methods that result in enumerations 

by surface area. The most common diversity 

index used for water quality studies is the Shannon, or Shannon-Wiener fndex (Shannon and 

Weaver, 1949) as shown below: 

where: 

ni = Total number of individuals in the 

i* taxon 

n= Total number of individuals 

s = Total number of taxa. 

(Washington (1984) provides a good explanation 

of how the index derived the name ShannonWiener Index rather than Shannon-Weaver Index.) 

theoretically, higher community diversity indicates better water quality (Wilhm, 1970). However, low diversity may be caused by factors other 

than water quality impacts, such as extremes in 

weather (floods or droughts), poor habitat, or 

seasonal fluctuations. Although diversity indexes 

such as the Shannon-Wiener Index still remain in 

widespread use (Washington, 1984), their limitations in accurately addressing a variety of perturbations has decreased their reliability (Cooke, 

1976; Hilsenhoff, 1977; Hughs, 1978; Chadwick 

and Canton, 1984; Washington, 1984; Mason et 

al., 1985; Resh, 1988). Kaesler er al., (1978) 

demonstrated that the popular Shannon’s Index 

was actuaIly not the preferred index for aquatic 

ecology studies and recommended the use of 

Brillouin’s (1962) Index. Resh (1988) reported 

that diversity indexes showed varied results in de8-10 
8-Freshwtater Benthi: M~t~~invertcbratc Community Structure and Function 

teding changes in water quality and that they are 

not the optimal measures of water quality. However, diversity indexes can provide additional 

information as to the community composition and 

should be reported if the data are available. Reliance on these indexes as the only, or predominant, 

measure on which water pollution control decisions are based is not valid. Washington (1984) 

provides an outstanding review of the history and 

uses of diversity indexes. 

Biotic Indexes-Biotic indexes use pollution 

tolerance scores for each taxon, weighted by the 

number of individuals assigned to each tolerance 

value. If desired, relative abundance measures can 

be used in biotic indexes. An example of a 

widely used biotic index (Hilsenhoff, 1977, 1982) 

is as follows: 

where: 

n, = Number of individuals in taxon i 

a, = Tolerance value assigned to taxon 

i 

n = Total number of individuals in the 

sample. 

Tolerance values can be found in Hilsenhoff 

(1987) or can be generated by regional-specific 

knowledge of the organisms’ tolerances. Typical 

ranges of organism index values are O-5,0-10, or 

O-11, with the higher numbers indicating greater 

tolerance to pollutants. Community indexes are 

generally limited to lotic systems impacted by 

organic enrichment (Woodiwiss, 1964, Chandler, 

1970; Hilsenhoff, 1977; Murphy, 1978; DePauw 

el al., 1986) or other general perturbations 

(Hawkes, 1979). Biotic indexes based on a 

specific population, rather than community, are 

addressed in the “Indicator Organisms” discussion 

below. Although the first widely appIied biotic 

index in this country was developed by Beck 

(1955) for Florida streams, the Hilsenhoff Biotic 

Index (Hilsenhoff, 1977, 1982) has gained great 

popularity and has been updated to revise the 

scoring system from a range of O-5 to O-11 (Hilsenhoff, 1987) and to include a family-level biotic 

index (Hilsenhoff, 1988). Because the biotic 

indexes rely heavily on known pollutiontoleranczs 

of the taxa, Washington (1984), Mason et ol. 

(1985), and Hawkes (1979) preferred the biotic 

indexes over the diversity indexes for water 

quality assessments. The success of the Hilstnhoff Biotic Index prompted use of the index, or 

modifications of it, in several state programs (e.g., 

Wisconsin, lllinois, New York, North Carolina) 

and EPA (Plafkin dr al., 1989) programs. Unfortunately, tolerance values are not available for 

many taxa because they tend not to exhibit water 

quality preferences, and the assessments are 

generally limited to organic enrichment. Washington (1984) provides an outstanding review of 

the history and uses of these indexes. 

Indkator Organisms--Indicator organisms have 

played a key role in the development of biotic 

indexes for both lotic and lentic systems. One of 

the first classifications based on indicator organisms was done in the lllinois River by Richardson 

(1928). Simpson and Bode (1980), Bode and 

Simpson (1982), and Rae (1989), among many 

others, used Chironomidae as indicator organisms 

for a variety of toxicants in stream systems. 

Hawkes (1979) provides an excellent review of 

the use of benthic macroinvertebrates for stream 

quality assessments, and Wiederholm (1980) does 

the same for lake systems. Data analyses for 

benthic macroinvertebrates in lentic systems have 

not been as progressive as those in lotic systems 

with regard to composite indexes and have relied 

extensively. on enumerations, diversity indexes, 

richness values, and indicator organisms (Fitchko, 

1986). Howmiller and Scott (1977), Krieger 

(1984), and Lauritsen et al. (1985) used oligochaete communities to establish a Great Iakes 

trophic index. Lafont (1984) also used oligochaetes to indicate fme sediment pollution. 

Briirkhurst et al. (1%8) and Winnell and White 

(1985) used chironomids to develop a similar 

index for the Great Lakes, and Courtetnan& 

(1987) classified Maine lakes using chironanid 

larvae similar to the studies of Saether (1979) and 

8-11 
Sediment Classification Methods Compendium 

Aagaard (1986) in European lakes. Hart and 

Fuller (1974) presented pollution ecology data for 

a number of freshwater benthic macroinvertebrates, as did U.S. EPA’s pollulion tolerance 

information series on Chironomidae (Beck, 1977), 

Trichoptera (caddisflies) (Harris and Lawrence, 

1978), Ephemeroptera (mayflies) (Hubbard and 

Peters, 1978), and Plecoptera (stoneflies) (Surdick 

and Gauh, 1978). Wasbington (1984) also 

reviewed population-based biotic indexes. 

Riaaess Measures-Richness measures are 

based on the presence or absence of selected taxa. 

Commonly used measures include the total number of taxa, the number of EPT (Ephemeroptera, 

Plecoptera, and Trichoptera), and the number of 

families. The higher the richness value is, the 

better the quality of the system. Richness measures have been shown to have low variability and 

high accuracy in identifying impact (Resh, 1988) 

and should be applied in each study. 

Enumerations-Enumerations involve obtaining 

a sample quantified by surface area to obtain 

specific abundances of each taxon. Examples 

include the number of total individuals, number of 

EPT individuals, ratios of number of individuals 

within a taxon to the total number of individuals 

(Ohio EPA, 1989a; Resb, 1988), and ratios of the 

number of individuals within one taxonomic group 

(e.g., EPT) to the number of individuals within 

another taxonomic group (e.g., Chironomidae) 

(Plafkin et ol., 1989; Resh, 1988). Interpretation 

of the enumeration ratios can be difficult without 

prior validaGon. Most possible enumerations 

comparing individual taxa to the total number of 

individuals are done for many studies, although 

the results may not be presented. The percent 

contribution of the individuals within a taxon at a 

sample site can be compared with the percent 

contribution at the reference sites to detect a 

change in community structure. Resh (1988) 

concluded that the seven common enumerations he 

tested bad extremely high variability and unacceptably low accuracy in detecting various impacts, and he suggested that they are not as useful 

for detecting environmental change as richness 

measures or the family biolic index. Although the 

measures Resh (1988) used may not be optimal 

for widespread use, they may still provide insight 

into changes in the community structure. Ohio 

EPA (1989a) has successfully used enumerations 

for the percentage of mayflies, caddisflies, Tanyiarsini midges, tolerant organisms, and “other” 

dipterans combined with non-insec! individuals as 

a basis for their stale biocriteria. 

Similarity Indexdrnmunity similarity indexes measure the similarity between benthic cornmunities at a reference and a study site, with high 

similarity indicating little change, or impact, 

between the two sites. The use of similarity 

indexes has been reviewed by Brock (1977) and 

Washington (1984). The simplest indexes to 

apply are ibose which use only the types of taxa 

found, not the abundance of the organisms within 

each taxon. The Jaccard Index (1908) and Van 

Horn’s Index (1950) are examples of the simpler 

indexes. Van Horn’s Index, used by Ohio EPA 

(1989b), is as follows: 

2W 

where: 

a = Number of taxa collected at one 

site 

b = Number of taxa collected at the 

olher site 

w = Number of taxa common to both 

stations. 

A value over 6.5 or 7.0 indicates good similarity. 

Plafkin et al. (1989) use the Jaccard Index in the 

rapid bioassessment protocols (RBPs). Other 

indexes such as the percent similarity (Brock 

1977) and the Bray-Curtis (1957) use the abundance of organisms. 

Functional Information-Community function 

measurements based on habitat, trophic structure, 

and other ecological measures were described by 

Kaesler et al. (1978) and used by Rooke and 

8-12 
9Freshwata Bent& Mt~~-~brvcrtebr~tc Commurity Structure and Function 

Mackie (1982a) as the “ecological community 

analysis” (EGA). Rooke and Mackie (1982b) 

reported the EGA to provide more information on 

environmental quality than diversity or biotic 

indexes, but the EGA was very time-consuming 

and not practical for rapid assessments. However, 

Cummins and Wilzbach (1985) and Cummins 

(1988) descrbe a rapid assessment method based 

on sampling coarse particulate organic matter and 

determining the functional feeding groups described in Merritt and Cummins (1984). This 

method is recommended in EPA’s RBPs (Plafkin 

et al., 1989). Rabeni er al. <1985) also described 

the usefulness of a functional feeding group 

approach to provide a “more ecologically sound 

classification of water quality” during their development of a biotic index for paper mill impacts. 

Another useful measure of function is observations 

of the incidence of morphological deformities in 

benthic macroinvertebrates, similar to the observations made for Karr’s Index of Biotic Integrity 

(IBI) for fish (Karr et al., 1986). Deformities 

have been associated with exposure of metals and 

organic compounds to Chironomidae (Cushman, 

1984; Cushman and Goyert, 1984; Wiederholm, 

1984b; Warwick, 1985; Warwick et al., 1987) and 

Trichoptera (Simpson, 1980; Petersen and Petersen, 1983). Karr and Kerans (1992) are developing an invertebrate IBI and have evaluated 10 

trophic and functional feeding group metrics. 

This promising work is continuing. 

Statistical Approaches-Various statistical 

approaches have been applied to determine whether the benthic community at a study site varies 

from that at a reference or other site. An excellent discussion of this issue appears in Klemm et 

~1. (1990) and USEPA (1992b). Depending on 

the chosen endpoints of the study, rigorous statistical analysis may not be necessary. For instance, 

if the endpoint is the number of taxa or richness 

measures, the variability is generally quite low and 

accuracy quite high. In this case, the differences 

between two communities would need to be 

evaluated based on study objectives. A “statistical” difference between two communities will not 

always indicate whether more subtle changes in 

community composition are occurring or whether 

mitigation may be warranted before a statistical 

change occurs. Sometimes when that change 

occurs, it is too late to protect the community. 

USEPA (1992b) has an outstanding discussion on 

applying uncertainty to decision-making. The 

same data evaluation procedures apply to both the 

marine and freshwater systems. The reader is 

referred to the statistical discussion in Chapter 9 

(marine bentbic community structure). 

Bivariate and multivariate analysis are often 

applied in benthic studies to define relationships 

between and among variables. Examples of these 

analyses include analysis of variance (ANOVA), 

correlations, regressions (including multiple 

regressions), and the two-sample t-test. A major 

drawback to these methods is the assumption that 

the data follow a statistical distribution such as a 

normal or lognormal distribution. This assump 

tion is often invalid when dealing with biological 

populations and communities. 

Alternatively, nonparametric analyses may be 

conducted. Such analyses are not based on assumptions about a specific distribution of the data. 

Examples of such tests include the chi-square test, 

binomial tests, rank correlations, or tests comparable to the t-test such as the Mann-Whitney test. 

Whichever statistical methods are employed, all 

data assumptions must be clearly stated and 

objectives known. 

8.2.1.2.4 Necessary Hardware and Ski& 

The hardware needed for field collection 

includes samplers (e.g., dredges, dip-nets), sieves, 

benthic macroinvertebrate containers, forceps, 

white enamel pans, ethanol preservative, and 

appropriate personal gear (e.g., hip boots or chestwaders, life vest if needed, and first aid kits). For 

the laboratory, standard biological laboratory 

equipment should be available, such as microscopes (both dissecting and compound), forceps, 

microscope slides and cover slips, ethanol, potassium hydroxide, mounting media, and sieves. A 

personal computer (containing a 2O-MB or larger 

hard drive) is important for storing and analyzing 

the data. 

Trained bentbic macroinvertebrate field biologists and taxonomists are needed for benthic 

8-13 
Sahent Classijication Methods Compendium 

community assessments. At least one should be 

proficient at identifications beyond the family 

level. That taxonomist should remain involved 

until the proficiency of the identifier in reaching 

family-level identifications is ensured. A minimum of a Master of Science degree in a related 

discipline is usually required for the taxonomist to 

have learned the necessary skills. However, 

adequate training is commonly available through 

taxonomy courses and workshops that can provide 

the necessary proficiency without an advanced 

degree. A demonstration of proficiency by accurately identifying a check sample prepared by U.S. 

EPA or a state agency is important. A trained 

benthic ecologist is necessary to compile and 

interpret the data. Although it would be ideal if 

the benthic ecologist had a rigorous statistical 

background, consultation with a statistician should 

be adequate. 

8.2.1.3 Adiyacy of Documentation 

There is ample documentation of both field 

methods and analytical techniques. The Journul of 

the North American 23enthoiogical Society is a prime 

w;rce of this information, as is technical exchange 

at professional meetings. Furthermore, there is a 

large volume of published and unpublished material 

that documents use of this method (USEPA 1992d, 

199le, 199Og, 1989f, 1988a). 

8.22 Applicability of Method b Human 

Health, Aquatic Life, or Wildufe 

R-&CtlOIl 

This method is directly applicable to the protection of aquatic life since it is based on direct measurements of benthic macroinvertebrates. This 

method is directly applicable to the protection of 

those aquatic organisms (e.g., fwh) and wildlife that 

directly feed on benthic macroinvertebrates (e.g., 

small mammals and wading shorebirds). It is 

indirectly appIicable to other wildlife that depend on 

benthos at other levels in the food chain. This 

method is also indirectly applicable to the protection 

of human health since benthic mauoinverlebrates 

tin serve as indicators of toxicant impacts that may 

affect humans via bioaccumulation pathways 

8.2.3 Abllity of Method to Generate 

Numerical Criteria for Spcclfk 

Cbcmlcds 

This method is used in conjunction with 

sediment toxicity and chemistry data to characterize toxicant impacts and assist with determining the appropriate levels at which the toxicants 

should be controlled. By itself, however, this 

method would not be used to generate chemicalspecific criteria. 

8.3 USEFULNESS 

8.3.1 Environmental Applkabilfty 

Benthic macroinvertebrates have been routinely used to assess environmental quality in a 

variety of geographical areas and ecoregions, as 

was discussed in Section 8.1. 

8.3.1.1 Suitability for Different Sediment Types 

Assessment of the freshwater benthic macroinvertebrate community structure is well suited 

for evaluating different sediment types since the 

benthos inhabit all substrates (Merrit and Cummins, 1984). Comparisons should be made 

among benthic communities of similar substrate 

since different types and numbers of organisms 

will inhabit different types of substrates. 

8.3.1.2 Suitability far Diflerent Chemicals or 

Classes of Chemicals 

Benthic macroinvertebrate communities are 

routinely used to assess potential impacts caused 

by many different chemicals or classes of chemicals. In addition to the uses described in Section 8.1.1.1 of this chapter, many benthic organisms are used to indicate stresses from specific 

chemicals or classes of chemicals (Brinkhurst et 

al., 1968; Hart and Fuller, 1974; Saether, 1979; 

Simpson and Bode, 1980; Wiederholm, 1980; 

Bode and Simpson, unpublished; Winnell and 

White, 1985; Aagaard, 1986; and Fitchko, 

1986). 

8-14 
g-Freshwater Bent?& Macr&wertebrak Community Structure and Function 

8.3.1.3 Suitability for Predicting Efiects on 

Diflerent Organisms 

The use of benthic macroinvertebrates as 

indicator organisms has already been discussed. 

Benthic macroinvertebrates can be used to predict the effects on other aquatic organisms because if the benthic macroinvertebrate community is impacted, then the impact is likely to be, or 

already has been, detrimental to other organisms. 

8.3.1.4 Suitability for In-Place Pollutant Control 

Benthic macroinvertebrates will best indicate 

in-place pollutant control needs through a sitespecific knowledge of surface water quality, 

habitat quality, and sediment chemistry and toxicity. Alone, the benthic macroinvertebrates can be 

used to screen for potential sources of sediment 

contamination based on spatial gradients in cornmunity structure, but they should not be used 

alone to definitively determine sediment quality or 

to develop chemical-specific guidelines. The 

benthic data must be integrated with other available data to determine sediment quality using a 

“weight-of-evidence” approach. 

831.5 Suitability for Source Control 

Benthic macroinvertebrates have been extensively used for source characterization and control in 

many of the state and U.S. EPA monitoring programs involving spatial surveys upstream and downstream of suspected sources (Ohio EPA, 1987; Bode 

and Novak, 1988; Courtemanch and Davies, 1988; 

Fiske, 1988; Maret, 1988; Penrose and Ovexton, 

1988; Shakelford, 1988; USEPA, 1991c, 1988a, 

198Sb; Fandrei, 1989). If a detrimental change is 

detected in the bentbic macroinvertebrate community 

and that change can be attriiutable to a source, then 

control measures can be implemented through the 

NPDES permit program. Many states aggressively 

pursue this action. 

8.3.1.6 Sui&bifi& for Disposal Applications 

The discussion presented in Section 9.3.1.6 of 

Qapter 9 (marine benthic macroinvertebrate community structure) is applicable to fresh water. 

Recenrly bent& community assessments have been 

required by U.S. EPA Region V, as stated in the 

Draft In&rim Guidance for the Design and Zhxubon of Sediment Sampling E~~oH.T Relating to Navigational Maintenance Dredging in Region V - Mhy 

1989 (USEPA, MM). In this guidance, be&tic 

lllX7OiIlVertebrate assessments are advised for areas 

that are suitable for open-lake disposal or for sediments that are difficult to characterize. All benthic 

community assessments will be made in concert with 

sediment chemistry and toxicity evaluations. 

83.2 Gcnerpl Advantages and Umttatbns 

The advantage of using the benthic macroinvertebrates community assessment approach to 

determining sediment quality is that it provides an 

economical and accurate indication of the health of 

the system under study, and it is based at dire& 

obsetvation rather than theoretically derived data. 

The major limitation is the difficulty in relating the 

findings to the presence of individual chemicak and 

specific concentrations of those chemicals for numeric in-place pollutant management. This method 

should be integrated with sediment chemistry and 

toxicity information. 

83.2.1 Ease of Use 

The equipment requirements for benthic surveys 

is minimal aad inexpensive compared to those for 

chemical/physical analyses or even toxicity tests. 

The organisms are easy to obtain, but difficult to sort 

and identity. All materials needed for benthic 

assessments are easily obtained through chemical and 

biological supply a>mpanies and require no special 

mechanical setup or calibration. 

83.2.2 Relative Cost 

The cost for benthic maaoinvertebrate assessments is economical compared to that for chemistry or toxicological evaluations. Ohio EPA 

(1990a) provided a cost of about $700 to conduu 

a benthic assessment at one sample site. However, this cost included overhead (e.g., rent, office 

equipment), all travel expenses, time spent in the 

8-M 
Saiiment Classifiuation M&ads Compendium 

field, and report preparation. Ohio EPA conducts 

artificial substrate (composite of five substrates) 

sampling in addition to natural substrate (multihabitat) sampling at each site. Their cost of 

$1,099 ($824 for artificial substrates and $275 for 

qualitative samples) was quite economical compared to chemical/physical testing ($1,653) or 

bioassay testing ($3,000 to $12,000) for each site. 

Plafkin et UL (1989) discussed staff requirements 

for sample collection and analysis. 

The most expensive items are the samplers 

and the microscopes to identify the organisms. 

However, most state programs and contractors 

have this equipment available for other program 

needs. The fieldwork can be conducted during the 

time it takes to collect a sediment sample. ‘Ihe 

most time-consuming aspect is the laboratory 

sorting and identifications, which may average 11 

hours per site. However, this process compares 

favorably with the amount of time required to set 

up and run a toxicity test or to prepare and analyze chemical variables. 

8.3.23 Tendency to Be Conservative 

The benthic macroinvertebrate community 

assessment provides a conservative measure, since 

the community is responding to both temporal and 

spatial perturbations. There are few chances, if any, 

of obtaining a result indicating a high-quality community when an impact occurs. Because of influences other than sedimentMater quality, it is more 

common to observe an impacted community when 

there is no sediment/water quality impact. Although 

the primary focus is on community-level information, changes in individual populations could also 

be addressed. However, the ecological significance 

of population changes may not be evident until the 

community is affected. 

In a review of surface water chemistry and 

benthic macroinvertebrate community assessments 

over t300 water body segment sites in Ohio, bioaiteria based on benthic macroinvertebrates were more 

sensitive (conservative) indicators of water quality 

(Ohio EPA, 199Ob). In 495 percent of the segments, the benthic and frsb assessment revealed 

impacts not detected by chemical water quality 

standards violations. In 47.4 percent of the sites, the 

chemical and biological assessment supporte4I one 

another. Only 2.8 percent of the sites did not have 

a biological impact when the chemistry indicated 

that there would be one. 

8.3.2.4 Level of Acceptance 

Benthic macroinvertebrate community asseSme&s of sediment,&vater quality have been used in 

freshwater systems since the early 1900s 

(Richardson, 1928). Most of the methods employed 

today have been widely accepti for use, although 

the use of function measurements is not as well 

documented. Perhaps the single most important 

demonstration of the level of acceptance of benthic 

assessments is the growing regulatory use and establishment of numerical biological criteria in state 

water quality standards. 

8.3.2.5 Ability to Be tmplemented by L-aboratoties 

with Qpical Equipment and Handhg 

F&li&kS 

The only special pieces of equipment required 

are the samplers and sieves, which are easily obtained from biological supply warehouses. Most 

biological laboratories will have disseding and 

compound micro~~pes, chemical reagents, microscope slides and cover slips, forceps, and any other 

materials needed. The laboratory’s capability to 

identify benthic macroinvertebrates is less common. 

Taxonomy is not a widespread skill and is more 

likely to be found in consulting fums than in analytical laboratories. 

8.3.2.6 Level of Eflort Required to Generate 

Resulti 

Depending on the study objectives and level 

of effort needed, results can be generated in 

written form in as little as 1 day (Plafkin et al. 

1989) or in several months. For example, Ohio 

EPA processes over 500 individual benthic samples each year, identifies the organisms, and 

prepares reports for regulatory use in less than 1 

year, with fewer than three full-time employees in 

their benthic macroinvertebrate unit. The critical 

period is the turnaround time for the taxonomy. 

8-16 
PFreshwater Benthic Macroinvertebrate Community Structure and Function 

With artificial substrates, an additional 6-week 

colonization period is required, unless a rapid 

assessment or moderate sized study is done, a 

written report including interpretation of results 

will typically require between 6 months and 1 

Year8.3.2.7 Degree to Which Results Lend 

Themselves to Interpretatiun 

It is never advisable to have an individual 

without training in benthic ecology interpret benthic 

data. Once the benthic ecologist provides a repoxt 

with recommendations, the results can be easily 

implemented into a management strategy. Although several numerical indexes that appear simple 

to use are available, data interpretation relies on all 

of the information generated for a study, including 

chemical, physical, and toxicological measurements, 

as well as indicator organisms and function measures. 

8.3.2.8 Degree of Environmental Applicability 

Benthic macroinvertebrate community strudure 

and function is used extensively to evaluate sediment and water quality and characterize impacts in 

lotic and lentic freshwater ecosystems. 

8.3.2.9 Degree of Accuracy and Precision 

Since benthic macroinvertebrates are measured 

directly, this method is highly accurate for characterizing sediment&vater quality effects on aquatic 

life. There is little chance, if any, that a highquality community will be indicated when an 

impact actually occurs (Type II error with a null 

hypothesis of no community change). Because of 

influences other than sedimentiater quality, it is 

more common to observe an impacted community 

when there is no sediment/quality impact (Type I 

error .with a null hypothesis of no community 

change). For environmental pollution control, a 

Type II error is much more serious than a Type I 

error, which is conservative. To reduce the possibility of a Type II error, the data (including chemistry and toxicity) must be interpreted by a trained 

benthic ecologist. Resh (1988) and USEPA 

(1992b) reviewed the levels of accuracy and precision for several of the data analysis techniques. 

To ensure as much accuracy and precision in 

the data as possible, a detailed Quality Assurance 

Program Plan should be established and followed. 

Careful and consistent field and laboratory v 

colsarenecessary. Itisalsomceswytosample 

during optimal conditions, which can minimize the 

effects of natural variations in the data. However, 

the natural variability, especially seasonal, is reduced when using a community-level approa& 

rather than a population-level approach. 

8.4 STATUS 

Sections 8.1.1 (Current Uses) and 8.3 (Usefulness) desaii the status of the discipline. 

8.4.1 Extent of Use 

?his method is widely used in both regulatory 

and nonregulatory sediment and water quality 

programs. It has been used to assess impacts due 

to organic enrichment and a variety of chemical 

classes in both lotic and lentic systems. Benthic 

macroinvertebrate community assessments are the 

most widely used instream biological measures in 

state water quality programs. 

8.4.2 Extent to Which Approach Has been 

Field-Validated 

Since it is an in siru study, field validation 

occurs when the approach can consistently and 

accurately assess environmental quality. Most 

benthic studies employ reference stations and rely 

on other environmental data to validate the method. 

The documentation provided in this paper should 

present adequate documentation of the method’s 

validity. 

8.43 Reasons for Ilmited use 

Benthic macroinvertebrate community assessments are very common in freshwater systems 

because of their relatively low cost and high information output. 

8-17 
Sediment Class$c&m Methods Comptndium 

8.4.4 Outlook for Future Use sad Amount 

of Jkvdopmeat Yet Needed 

The outlook for the future use of benthic 

macroinvertebrate community structure and function in sediment quality assessment is very good 

because of the recognition that benthic macroinvertebrates provide substantial information that 

the chemistry and toxicity data alone cannot 

provide. With the Clean Water Act mandate to 

maintain and restore biological integrity, benthic 

community assessments can help determine whether sediment quality is impairing the designated 

uses and biotic integrity. With the increasing 

reliance on numerical biocriteria, additional sediment quality problems will be identified. The area 

where development is most needed is in combining benthic community assessments with chemical 

and toxicological data in an integrated approach 

for assessing sediment quality. In addition, the 

functional measures, which also hold much promise for sediment assessments, need to be validated 

more thoroughly. 

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Yasuno, M., Y. Sugaya, and T. Iwakuma. 1985. 

Effeds of insecticides on the benthic cornmunity in a model stream. Environ. Poilut (Ser. A) 

38:31-43. 

8-26 
CHAPTER 9 

Marine Benthic Community Structure 

Assessment 

Betsy Striplin, Gary Braun, and Gordon Bilyard 

Tetra Tech, Inc. 

11820 Northup Way, Suite 100E, Bellevue, WA 98005 

(206) 822-9596 

Benthic communities are communities of 

organisms that live in or on the sediment. In most 

benthic community structure assessments, primary 

emphasis is placed on determining the species that 

are present and the distribution of individuals 

among those species. These community attributes 

are emphasized largely for pragmatic reasons. 

Although it is relatively simple to collect, identify, 

and enumerate benthic organisms, it is very 

difficult to determine first-hand the spatial distributions of species and individuals within the 

benthic habitat, or the functional interactions that 

occur among the resident organisms or between 

the resident organisms and the abiotic habitat. 

Hence, information on benthic community composition and abundance is typically used in conjunction with information in the scientific literature to 

infer the distributions of species and individuals in 

three-dimensional space and the functional attributes of the community. Because ail of the major 

structural and functional attributes of benthic 

communities are affected by sediment quality in 

generally predictable ways, benthic community 

structure assessment isa valuable tool for evaluating sediment quality and its effects on a major 

biological component of marine, estuarine, and 

freshwater ecosystems. 

Benthic habitats may be broadly divided into 

hard-bottom habitats and soft-bottom habitats. 

Many types of each exist in marine, estuarine, and 

freshwater ecosystems. Hard-bottom habitats 

include rocky shorelines and bottoms of lentic and 

lotic systems, rocky intertidal and subtidal habitats 

in marine and estuarine systems, and coral reefs. 

Soft-bottom habitats include mud and sand habitats in marine, estuarine, and freshwater systems; 

marine, estuarine, and freshwater macrophyte 

beds; freshwater wetlands; and estuarine salt 

marshes. Each of these habitats requires different 

sample collection methods and different survey 

design considerations. The emphasis of this 

chapter is on assessments of marine benthic 

community structure in soft-bottom habitats as an 

indicator of sediment quality. Freshwater benthic 

invertebrate community structure is discussed in 

Chapter 8. 

9.1 SPECIFIC APPLICATIONS 

Assessment of benthic community structure is 

an in situ method that can be used alone, as part 

of other approaches [e.g., Sediment Quality Triad 

(see Chapter 10) and Apparent Effects Threshold 

(AET) (see Chapter 11)], or in combination with 

other sediment assessment techniques (e.g., sediment toxicity bioassays). It is commonly used in 

three ways to assess impacts to benthic communities and sediment quality: 

• To compare test and reference stations, 

for the purpose of determining the spatial 

extent and magnitude of such impacts; 

• To identify spatial gradients of impacts; 

and 

• To identify temporal trends at the same 

locations through time. 

By definition, benthic communities include all 

organisms living on or in the bottom substrate. 

For practical reasons, assessments of benthic 

community structure in soft sediments usually rely 

on the macrofauna (i.e., organisms retained on a 

l.0- or 0.5-mm sieve) and to a lesser extent the 

meiofauna (i.e., multicellular organisms that pass 
Sediment Classification Methods Compendium 

through a l.0- or 0.5-mm sieve). Reasons for the 

more limited use of meiofauna are twofold: 

• Although they may be sampled quantitatively, their small size makes working 

with them difficult, and the taxonomy of 

many of the groups (e.g., nematodes) is 

not well known. 

• The functional attributes of the various 

meiofaunal taxa are poorly known, and it 

is therefore difficult to interpret the importance of the presence or absence of the 

various taxa in relation to environmental 

quality. (For example, knowledge of 

meiofaunal taxa that respond positively or 

negatively to organic enrichment of the 

sediments is extremely limited.) 

Difficulties in quantitatively sampling other size 

classes of benthic organisms such as the megafauna (i.e., large organisms that are typically 

measured in centimeters) and the microfauna (i.e., 

microbes) usually preclude them from consideration in assessments ofbenthic community structure. Furthermore, although the functional importance of sediment microbes has been studied, their 

structural and functional characteristics have not 

been used as indicators of sediment quality. 

9.1.1 Current Use 

Assessments of benthic community structure 

have been used to describe reference conditions, 

baseline conditions, and the effects of natural and 

anthropogenic disturbances. Selected examples of 

current uses of this approach are provided below. 

Organic Enrichment-Pearson and Rosenberg (1978) performed an extensive review of 

benthic community succession in relation to 

organic enrichment of marine and estuarine sediments. Based on that review, they developed a 

generalized model of structural community changes (i.e., numbers of species, abundances, biomass) 

in relation to organic enrichment, and identified 

opportunistic and pollution-tolerant species that 

are indicative of organic enrichment. Concepts 

developed by Pearson and Rosenberg (1978) have 

subsequently been used by many investigators to 

assess the degree of organic enrichment that has 

occurred in a variety of soft-bottom habitats. For 

example, Dauer and Conner (1980) assessed the 

effects of sewage inputs on benthic polychaete 

populations in a Florida estuary by collecting 

information on the total number of individuals, 

total biomass, and average number of species. 

They compared the sewage-affected site with a 

reference site and examined the response of 

individual species to organic enrichment. In 

another study in Florida, Grizzle (1984) identified 

indicator species based on life history responses to 

organic enrichment and other physicochemical 

changes. The taxa identified as indicator species 

in enriched areas were generally characterized by 

opportunistic life history strategies. Vidakovic 

(1983) assessed the influence of domestic sewage 

on the density and distribution of meiofauna in the 

Northern Adriatic Sea. He concluded that raw 

domestic sewage did not have a negative influence 

on the density and distribution of meiofauna, but 

the nematode/copepod ratio (Parker, 1975) indicated that these. stations were under stress. 

Contamination Due to Toxic Metals and 

Metalloids-Rygg (1985a, 1986) assessed benthic 

community structure in Norwegian fjords where 

the disposal of mine tailings had resulted in metals 

contamination of the sediment. His studies 

showed an inverse relationship between concentrations of metals in the sediment and the species 

richness and abundance of the benthic macroinvertebrate fauna. Bryan et al. (1987) examined 

population distributions of the oyster Ostrea 

edulisy, the polychaete Nereis diversicolor, and 

the cockle Cerastoderma edule in relation to 

wastes from metals mining in the Fal Estuary. 

They concluded that the distribution of species is 

dependent on their ability to tolerate copper and 

zinc, and on the capabilities of a population to 

develop a resistance to metals and thereby maintain their original distribution range. 

Contamination Due to Toxic Organic Compounds-Toxic organic compounds are frequently 

associated with municipal discharges, industrial 

effluents, and storm drains. These discharges may 

also result in organic enrichment and contamination by metals or metalloids. The following 

benthic studies provided evaluations of sediment 

9-2 
94arinc Benthic Community Structure Assesmmt 

quality in areas primarily affected by toxic organic 

compounds: 

n Czeosote contamination. Tagatz et ol. 

(1983) examined the benthic communities 

that colonized uncontaminated sediments 

and sediments contaminated with three 

different concentrations of creosote (177, 

844, and 4,428 pg/g) in field and laboratory aquaria to assess the effects of 

marine-grade creosote on community 

structure. Numbers of individuals and 

numbers of species in field-colonized 

communities were significantly lower in 

all three creosote-contaminated sediments 

than in the controls. In the laboratorycolonized communities only the two 

higher creosote concentrations had reduced numbers of individuals and species. 

Distribution of individuals within species 

was similar for the laboratory and field 

assemblages of animals. 

n Oil contamination. Elmgren et of. (1983) 

determined that acute effects of the Tsesir 

oil spill were noted after 16 days on both 

the macrofauna and meiofauna. Initial 

recovery was noted 2 yr after the spill. 

However, the authors predicted that complete recovery would require at least 5 yr. 

Jackson et al. (1989) investigated the effects of spilled oil on the Panamanian 

coast and found that shallow subtidal reef 

corals and the infauna of seagrass beds 

had experienced extensive mortality. 

After 15 yr, only some of the organisms 

in areas exposed to the open sea had 

recovered. Clifton et al. (1984) performed field experiments in Willapa Bay, 

Washington, and found that oil in the 

sediments modified the burrowing behavior of infaunal benthos. 

Dredging and Construction-Related Activities-Swartz et al. (1980) examined species 

richness and species abundances just before 

dredging occurred in Yaquina Bay, Oregon, and 

for 2 yr after dredging. Benthic community 

recolonization was followed from the appearance 

of opportunistic taxa through their replacement by 

less tolerant taxa. Rhoads ef ol. (1978) examined 

the influence of dredge-spoil disposal on be&tic 

infaunal succession in Long Island Sound by 

classifying species into groups based on their appear8nce in a disturbed area. They suggested that 

the “equilibrium community is less productive 

than a pioneering stage” and suggested that productivity may be enhanced through managed 

disturballcts. The abundance of polychaetes, 

molluscs, and crustaceans is currently used to help 

assess potential biological effects of dredged 

material disposal by the Puget Sound Dredged 

Disposal Analysis Program (SAIC, 1991; Striplin 

et al., 1991). 

Natural Disturbances-Most stud&s of 

natural disturbances have assessed the recovery of 

benthic communities after the disturbance (e.g., 

large storms and associated wave activity, oxygen 

depletion, salinity reductions, El Nifio). For 

example, Dobbs and Vozarik (1983) sampled 

stations before and after Storm David and observed that the number of species decreased after 

the storm. They also documented changes in the 

rank order of the dominant taxa. Santos and 

Simon (1980) examined defaunation of benthic 

communities before, during, and after annual 

hypoxia in Biscayne Bay. They documented that 

recolonization occurs fairly rapidly after the 

defaunation period. Oscillations in macrobenthic 

populations in the shallow waters of the Peruvian 

coast were examined by Tarazona ef 41. (1988). 

Fluctuations in density, biomass, species composition, and diversity were attributed to the El Niiio 

of 1982-1983. 

Assessment of benthic community structure is 

also used as a component of other sediment 

quality assessment tools. Along with sediment 

chemistry and sediment toxicity bioassays, it is 

one of three components of the Sediment Quality 

Triad (see Chapter 10). It is also a component of 

the Apparent Effects Threshold approach (see 

Chapter 11). 

9.1.2 Potential Use 

To date, benthic community assessments 

performed to evaluate sediment quality have 

9-3 
Sediment Class~~tion Methods Compendium 

focused on the relationships between community 

variables (e.g., numbers of species, total abundance, biomass) and measures of sediment quality 

(e.g., organic content, concentrations of chemical 

contaminants). Only for organic enrichment have 

individual species been identified that are indicative of various degrees of sediment alteration [see 

for example Pearson and Rosenberg (1978), Word 

et al. (1977)J. Moreover, for only a very few 

species has the autecological relationship between 

organic enrichment of the sediments and an 

individual species been explored. [For example, 

Fabrikant (1984) explored the autecology of the 

bivalve mollusc Parvilucina tenuisculpta in relation to organic enrichment of the sediments in the 

Southern California Bight.] A tremendous potential exists, however, for identifying species that 

are indicative (by their persistence, enhanced 

abundance, reduced abundance, or absence) of 

sediment contaminants at various concentrations. 

The identification of such taxa will not be simple 

because of the complex ecological interactions that 

occur within benthic communities, and because 

sediments are frequently contaminated with a 

mixture of chemicals. A first step in this process 

might be to attempt to identify species or suites of 

species that could be used to separate the effects 

of sediment organic enrichment from sediment 

contamination by toxic substances. 

Another potential use of benthic community 

assessments would be to predict recovery of 

benthic habitats following the execution of remedial actions at contaminated sites. To date, it has 

not been possible to use extant benthic community 

structure to predict recovery because the only 

model that relates benthic community structure to 

sediment quality [i.e., the Pearson and Rosenberg 

(1978) model] is not quantitative. Quantification 

of this model and the development of quantitative 

models for other sediment contaminants will be 

required before benthic community assessments 

can be used to predict sediment quality. A valuable byproduct of such models would be the 

ability to predict the capacity of the remediated 

area to support higher trophic level organisms that 

forage on benthic organisms, including commercially and recreationally harvested demersal fishes. 

9.2 DESCRIPTION 

9.2.1 Deseriptioa of the Method 

An assessment of benthic community structure 

typically involves a field survey that includes 

replicated sampling at each station; sorting and 

identification of the organisms to species or lowest 

possible taxon; analyses of the numbers of taxa, 

numbers of individuals, and sometimes biomass in 

each sample; and identification of the dominant 

taxa. Results of the field survey are then interpreted in conjunction with other sediment variables (e.g., sediment grain size, total organic 

carbon) that were collected concurrently with the 

benthic samples. 

9.2.1. I Objectives and Assumptions 

The objective of the benthic community 

structure approach is to identify degraded and 

potentially degraded sediments by examining the 

communities of organisms that inhabit those 

sediments. This empirical approach assumes the 

following: 

n Because benthic infauna are generally 

sedentary, benthic community structure 

reflects the chemical and physical environment at the sampling location. 

n Benthic community structure may be 

altered in a predictable manner over time 

and space by chemical or physical disturbances. 

n The execution of proper data collection 

and analysis methods can reduce natural 

variability of benthic infaunal data and 

enable the detection of trends in sediment 

quality. 

9.2.1.2 Level 0fEflor-t 

The level of effort required to assess benthic 

community structure is relatively high. Regardless 

of the analytical methods, a field survey is required to collect the organisms. The sorting and 

identification process is IaborGntensive and usu94 
9-Marine Ben thic Community Structure Assessment 

ally expensive. Program objectives will determine 

whether the data analyses are simple or complex. 

9.2.1.2.1 Type of Sampling Required 

The type of sampling required to tolled benthic 

organisms is dependent on the objedives of the 

sampling program and on the area under study. 

Usually, the objective of a benthic sampling 

program is to study the characteristics of and the 

variation in the benthic community that occupies 

specific sampling stations. In this case, all organisms present in the sediment at that location are 

sampled together: those that normally reside in 

the surface few centimeters of sediment and those 

that normally reside deeper in the sediment (e.g., 

S-15 cm below the surface). In some instances, a 

sampling program may have a different objective. 

For example, sampling for the Benthic Resources 

Analysis Technique (BRAT) (Lunz and Kendall, 

1982) involves collecting box core samples and 

determining the biomass (and possibly the communities) present in specific sediment strata (i.e., 

O-2 cm, 2-5 cm, 5-10 cm, and lo-15 cm below the 

sediment surface). In that technique, the benthic 

data are compared with the benthic organisms 

consumed by bottom-dwelling fish (as determined 

by gut content analyses of fish captured in the 

same area) to determine the food value of the 

benthos. 

Characteristics of the area under study also 

influence the type of sampling. In intertidal or 

littoral environments where sampling stations can 

be occupied by walking to the site, samples are 

usually collected using a hand-held corer. If 

stations are located in subtidal areas, then remote 

sampling from a vessel is performed using a box 

corer or grab sampler. Sediment grain size may 

influence final selection of the sampler. Some 

samplers (i.e., many box corers) perform poorly in 

sandy sediments, whereas others (i.e., van Veen 

grab, Smith-McIntyre grab) perform adequately in 

a greater range of sediment types (i.e., tine to 

medium sand, silt, silty clay). Methods and 

equipment for sampling infaunal communities are 

further described in several publications (Word, 

1976; Swartz, 1978; Eleftheriou and Holme, 1984; 

Nalepa et al., 1988). Blomqvist (1991) provides 

an extensive review of quantitative sampling 

methods, including a detailed bibliography of 

pertinent papers. 

Program objectives and knowledge of benthic 

communities in the study area will influence 

selection of the sieve size through which sediment 

samples will be washed. It is important that the 

sieve mesh size-s be appropriate for the community 

under study (e.g., 64 /an for meiofauna, 05 or 1.0 

mm for maaofauna). Generally, the chances of 

retaining most macrofauna species and individuals 

(and therefore increasing sampling accuracy) are 

improved by the use of a finer mesh (but, see 

Bishop and Hartley, 1986). However, sieve size 

is an important determinant of the cost and level 

of effort necessary to obtain quantitative data. 

Very little difference in the field processing time 

exists between use of a 0.5~mm and a 1.0~mm 

sieve when sieving sediments finer than coarse 

sand, but laboratory analyses are much more timeconsuming when the smaller mesh is used because 

it retains more abiotic materials and many smaller 

organisms. 

9.2.1.2.2 Methods 

Methods for collecting data on benthic community structure are divided into three categories: 

program design, field methods, and laboratory 

methods. Each of these categories is briefly 

discussed below. 

Program design includes the selection of 

station locations, level of replication, type of 

sampler, screen size, data analysis methods (discussed later), and quality assurance/quality control 

(QA/Qc) procedures. The selection of station 

locations will diredly influence the usefulness of 

the resulting data. Stations that will be compared 

to one another (including reference stations) 

should be situated in areas with similar hydrography, water depth, and grain size to minimize 

the natural variability in benthic community 

composition that can be attributed to these factors. 

However, such station placement is not always 

attainable because of altered grain size distributions that often result from contaminant sources. 

Selection of the number of replicates is an 

important component of program design because 

9-5 
Sediment Class;fic(lCiun Methods Compendium 

the accuracy and precision with which benthic 

community variables are estimated depend in part 

on the size of the sample (including all replicates). 

For example, the abundance of a single taxon is 

generally a less accurate descriptive variable than 

is the abundance of the total taxa because of the 

greater variability typically associated with one 

taxon in comparison with the sum of all taxa. 

The total area sampled among the replicates at 

each station should be large enough to estimate a 

given variable within the limits of accuracy and 

precision that are acceptable to meet study objectives. A single sample may be useful for general 

distributional or trends analyses (Cuff and Coleman, 1979), but the inherent patchiness of benthic 

communities makes collection of a sufficient 

number of replicate samples (a minimum of 3-5, 

depending on study objectives and sampler area) 

necessary to ensure statistical reliability (see 

Elliott, 1977). Within a study area, adequate 

sample size may be determined by maximizing the 

number of species collected or by minimizing the 

error associated with the mean for the variable in 

question (Gonor and Kemp, 1978). Additional 

research on replication is presently being conducted by EPA in Newport, Oregon, under the direction of S. Ferraro (Swartz, R.C., 15 March 1989, 

personal communication). 

Power analysis can assist in determining the 

appropriate number of replicates. A power analysis includes consideration of the minimum detectable difference in selected biological variables 

(i.e., the minimum difference in mean values of a 

variable at several stations that can be detected 

statistically, given a certain level of variability 

about those mean values) and the power of the 

statistical test to be used. The power of the test is 

especially important because it defines the probability of correctly detecting experimental effects 

(e.g., differences in biological variables among 

sampling stations). For a specified variance 

associated with a biological variable, the statistical 

power of a test and the minimum detectable 

difference among sampling areas can be expressed 

as a function of sample size. The allocation of 

sampling resources (stations, replication, and 

frequency) can then be determined with regard to 

available resources, practicality of desigrr, and 

desired sensitivity of the subsequent analyses. 

Discussions and examples of this approach are 

found in Winer (1971), Saila et al. (1976), Cohen 

(1977), Moore and McLaughlin (1978); Bros and 

&well (1987), Ferraro et aI. (1989), Kronberg 

(1987), Tetra Tech (1987), Self and Mauri&en 

(1988), and Vezina (1988). 

A potential drawback to use of power analysis 

is that it requires aprbri knowledge of variability 

in the benthic communities that will be studied. 

If such variability is not known and cannot be 

estimated, then the number of replicates will 

probably reflect either funding limitations or 

generally approved sampling methods. For example, Eleftheriou and Holme (1984) and Swartz 

(1978) recommend that an area of 0.5 m2 be 

sampled to assess species composition in coastal 

and estuarine regions. Most studies of benthic 

community structure routinely involve five replicate 0.1-m* grab samples. A single 0.1-m* grab 

sample may be sufficient to obtain “useful descrip 

tive information” for use in cluster analyses 

(Word, 1976). However, a single sample precludes direct estimates of within-group variance 

for statistical analyses. Because individuals are 

distributed logarithmically among the species of a 

benthic community (Preston, 1948; Sanders, 1968; 

Gray and Mirza, 1979), species collected in the 

second and successive replicates that were not 

collected in any of the previous replicates most 

often will be numerically “rare.” Note that “rare” 

is not synonymous with “unimportant.” Hence, a 

single 0.1-m’ sample is generally not adequate to 

characterize benthic community structure and 

function. In general, five 0.1-m’ grab samples are 

recommended for determining benthic community 

structure, unless evaluation of site-specific data 

(i.e., a power analysis) indicates that sufficient 

sensitivity can be obtained with fewer samples, or 

that a greater number is required due to extreme 

spatial heterogeneity. (Note that at least three 

samples are required for parametric statistical 

analyses.) 

Another aspect of program design is selection 

of the appropriate degree of navigational accuracy. 

For baseline or distributional studies, repeatable 

station location may not be a high priority, and 

methods such as Loran C may be sufficient. 

9-6 
9-Marine Benfhic Communify Stndure A-t 

However, for monitoring programs where reoccupation of exact stations is important (e.g., disposal 

site monitoring), a more accurate positioning 

method (e.g., an electronic distance-measuring 

device or Mini-Ranger) may be required. 

A quantitative sampling device and an appropriate mesh size must be selected to ensure that 

size classes of organisms appropriate for assessing 

sediment quality are collected. Selection of a 

sampler and sieve are discussed above, in Section 

9.2.1.2.1. 

Field and laboratory methods must be conducted according to rigorous QA/QC protocols. 

Field methods include collecting, sieving, and 

preserving the samples. Samples are typically 

preserved in a solution of 10 percent buffered 

formalin for at least 24 h. Laboratory methods 

include rinsing the formalin solution from the 

samples within 7-10 days, followed by storage in 

70 percent ethanol. Samples are sorted under a 

dissecting microscope during which all organisms 

are removed from the samples and placed in vials 

for identification and enumeration of individual 

taxa. ‘Ihe time required to sort and identify a 

benthic sample varies greatly depending on the 

sieve size, sample area, and sediment composition. 

Sorting may take as little as 1 h for a 0.1-m* 

sample sieved through a l.O-mm screen, or as 

much as 12 h if wood chips or other debris are 

present. The time needed to identify organisms in 

a sample depends on the number of organisms 

(which is a function of sieve size, habitat, or 

degree of contamination) and number of taxa 

present. The number of hours needed to identify 

organisms in a sample may range from 1 to over 

10 b. 

In addition to the collection of samples for 

analysis of benthic community structure, separate 

sediment samples should be collected at all stations for conventional sediment chemistry variables (e.g., sediment organic content, sediment 

grain ‘size distribution). Because organic carbon 

content and sediment grain size naturally affect 

the composition of benthic communities, measurement of these variables will assist in determining 

whether benthic communities are affected by 

reduced sediment quality. 

9.2.1.2.3 Types of Data Required 

The two primary structural attributes of any 

benthic community are the distribution of species 

and individuals in threedimensional space, and 

the distribution of individuals among species and 

higher taxa. Given an understanding of these two 

structural attributes, it is possible to infer ftmdional.attributes of the benthic community, including 

trophic relationships, primary and secondary 

productivity, and interactions between the resident 

bioia and the abiotic habitat. l’he data required 

for analysis of the structural and functional at&ii 

utes include the number of taxa (identifications 

should be to the lowest taxonomic level possible), 

the abundance of each taxon, biomass (depending 

on program objectives), and conventional sediment 

chemistry variables. However, collection of the 

appropriate data does not ensure proper evaluation 

of the structural and functional attributes. ‘Ihe 

selection and implementation of data analyses are 

equally important, and are discussed in the remainder of this section. The data analyses presented in this section address primarily structural 

components of benthic communities. However, 

functional attributes can be inferred from many of 

those structural attributes. 

Various types of data analyses are used to 

describe benthic community structure, depending 

on the objectives of the particular program 

However, several descriptive values are common 

to most program objectives. All organisms collected in each sample are enumerated (i.e., total 

abundance), and abundances of major taxonomic 

groups are usually summarized. Depending on the 

level of identification, abundances of individual 

taxa, numbers of taxa, and lists and abundances of 

pollution-tolerant and pollution-sensitive taxa in 

each sample may be developed. Biomass of major 

taxonomic groups and total biomass are sometimes 

reported. The composition of the numerically 

dominant taxa are analyzed when species level 

identifications are performed. In addition, desaiptive indexes such as diversity [the distribution of 

individuals among species; see Washington (1984) 

for additional definitions of diversity], eveuncgs 

(the evenness with which individuals are distributed among taxa), and dominance (the degree to 

9-7 
St&rrerrt Classijicntion Methods Compendium 

which one or a few species dominate the community) are usually calculated. 

Most programs evaluate the temporal or 

spatial differences in benthic community structure. 

Typically, comparisons of one or more indexes are 

made at the same station over time and compared 

to a baseline value, or comparisons are made 

between stations in a study area and stations in a 

reference area If an adequate number of samples 

is collected (i.e., three or more), statistical tests 

such as t-tests or Analysis of Variance (ANOVA) 

(or their nonparametric analogues) are often 

performed to determine whether significant spatial 

or temporal differences exist among benthic 

communities. 

Besides univariate (i.e., single-variable) statistical analyses, multivariate (i.e., multiple-variable) 

analyses are frequently performed (e.g., Boesch, 

1977; Green and Vascotto, 1978; Gauch, 1982; 

Shin, 1982; Long and Lewis, 1987; lbanez and 

Dauvin, 1988; Nemec and Brinkhurst, 1988a,b; 

Stephenson and Ma&e, 1988). Multivariate 

analyses include classification methods (i.e., 

grouping similar stations into clusters) and ordination methods [i.e., representing sample or species 

relationships as faithfully as possible in a lowdimensional (two-four dimensions) space]. [See 

Gauch (1982) for an overview of multivariate 

methods. ] Multivariate techniques group data and 

display them on a two-dimensional plot or dendrogram so that stations exhibiting similar communities are located closer to one another than to 

stations with dissimilar communities. The numerical and graphical results can then be compared 

with physical and chemical data collected concurrently to determine whether those variables 

correlate with trends in benthic communities. A 

commonly used classification technique involves 

first computing a matrix of similarity indexes that 

represent the degree of similarity in species composition between two stations. Commonly used 

similarity indexes include Bray-Curtis, Canberra 

metric, and Euclidian distance indexes. The 

similarity matrix is then entered into a clustering 

algorithm (e.g., pair-wise averaging, flexible 

sorting) to produce a dendrogram depicting similarities among stations. Commonly used ordination techniques include principal components 

analysis, detrended correspondence analysis, and 

discriminant function analysis. Bernstein and 

Smith (1986) developed an index of benthic 

community change along pollution gradients that 

is derived from results of ordination analysis. The 

index (called Index 5) is a measure of change 

from reference conditions. 

Benthic community surveys generate large 

data matrices. These data matrices are often 

reduced by the elimination of certain species 

(Boesch, 1977) prior to performing multivariate 

analyses. A variety of methods exist for reducing 

data matrices (see Stephenson el of., 1970, 1972, 

1974; Day et al., 1971; Clifford and Stephenson, 

1975). 

Both parametric statistical tests and multivariate analyses may involve data transformations. 

Transformations of the original data may be 

necessary for one or more of tbe following reasons: 

n Benthic data sets are usually characterized 

by large abundances of a few species and 

small abundances of many species; 

n The distrrbution of individuals among 

species tends to be lognormal; and 

n Sampling effort may be inconsistent 

(Boesch, 1977). 

The two basic types of transformations are strict 

transformations and standardizations. S&id 

transformations are alterations of the original 

values (e.g., species abundances) without reference 

to the range of values within the data. Commonly 

used transformations are square root, logarithmic, 

and arcsine (Sokal and Rohlf, 1981). Standardizations are alterations that depend on some poperty 

of the data under consideration. A common standardization is the conversion of values to percentages. 

Benthic data are transformed to better meet 

the assumptions of parametric tests (e.g., normality, homogeneity of variances). ln multivariate 

analyses, data are often transformed using logarithms [e.g., log (x+1)] because of the presence of 

zero scores. This transformation is also applied 

9-8 
9--Marine Benthic Community Structure Assessment 

when population variance estimates are positively 

correlated with mean values (Sokal and Rohlf, 

1981). Clifford and Stephenson (1975) discuss in 

detail the effects of transformations on commonly 

used resemblance measures. 

Benthic community structure is usually compared with chemical and physical data that are 

collected concurrently. These comparisons may 

take the form of simple linear correlations, correlations with cluster groups, or correlations using 

multivariate techniques such as disaiminant 

analyses. Multiple discriminant analysis attempts 

to isolate groups of similar stations so that viuiables responsible for the separation of groups can 

be identified. Results may be used to determine 

whether differences in community structure are 

due to variations in sediment grain size, variations 

in other physical characteristics of the environment, or changes in sediment quality due to toxic 

substances or organic materials. 

The use of different methods and analyses 

may result in different interpretations of the same 

data. For example, use of the same data with 

different standardization methods in a classification analysis can yield very different results (Austin and Grieg-Smith, 1968). Generally, the more 

analyses that are conducted on the data, the higher 

the probability of interpreting the data accurately. 

9.2.1.2.4 Necessary Hardware and Skills 

The hardware needed to perform a benthic 

community assessment is fairly common and 

should be readily available. Equipment includes 

field collection gear (e.g., sampling vessel, appropriate sampler, sieves, sample storage containers, 

buffered fixative) and standard biological laboratory equipment (e.g., microscopes, sieves, hydrometers or pipets, and a balance). More specialized 

equipment includes a muffle furnace for determining total volatile solids concentrations, a taxonomic reference collection, and a taxonomic reference 

library. Computer equipment and appropriate 

software are required to make studies costeffective. A microcomputer is sufficient for most 

analyses, but some complicated multivariate 

analyses may require the use of a minicomputer or 

mainframe computer. 

Trained benthic taxonomists are required to 

ensure accurate identifications. Some computer 

programming and some level of data management 

are usually required. A trained benthic ecologist 

is required to synthesize and interpret the data. 

However, the amount of training depends on the 

required level of interpretation. For example, 

interpretation of several multivariate methods 

would require a higher level of training than interpretation of descriptive indexes. 

9.2.1.3 A&quncy of Documentation 

Many different approaches and methods are 

used to analyze benthic data, some of which have 

their origins in classical terrestrial community 

ecology. Because analysis of benthic community 

structure is a relatively old assessment tool, literally thousands of papers have been written about 

the method. Several books and protocols have 

also been developed to describe field and laboratory techniques [e.g., Holme and McIntyre (1984), 

Puget Sound Protocols (Tetra Tech, 1986b), U.S. 

EPA 301(h) protocols (Tetra Tech, 1986a)]. 

However, a comprehensive document that describes standardized procedures for analyzing and 

interpreting benthic community data is lacking. 

The most commonly used interpretive approaches include measures of diversity and classification. Sometimes a general consensus exists 011 

the best techniques to use within an approach 

(e.g., widespread use of Shannon-Wiener diversity 

index, although there is debate as to whether this 

is a suitable index for environmental impad 

analysis). Despite this consensus, studies do not 

necessarily follow a specified format. Program 

objectives tend to dictate the types of hypotheses 

posed and analyses used. Many relatively new and 

exciting approaches have been proposed for 

assessing benthic community structure. However, 

most are relatively untested and are not widely 

used [e.g., benthic resource analysis technique 

(Latnz and Kendall, 1982), abundance-biomass 

comparison (Warwick 1986; Warwick et 01, 

1987), infaunal trophic index (Word, 1978,1980), 

aematode:copepod ratio (Amjad and Gray, 1983; 

Lambshead, 1984; Shielis and Anderson, 1985; 

Raffaelli, 1987), lognormal distribution (Gray and 

9-9 
Sediment Classificdion Methods Compendium 

Mirza, 1979), Index 5 (Bernstein and Smith, 

1986)]. Each of these methods has shown promise in some situations, but more testing and validation are needed before any can gain universal 

acceptance. 

Very few assessments of the information 

gained from analyses of data at the species level 

vs. the major taxa level have been undertaken. 

Warwick (1988) evaluated the results of ordinations run on various hierarchical levels of taxonomic data for five data sets. Three of the data 

sets were of macrofauna (from Loch Linne, Clyde 

Sea, and Bay of Morlaix); one was of nematodes 

from the Clyde Sea; and the last was of copepods 

from Oslofjord that were subjected to different 

levels of particulate organic material. He reported 

that in none of those five cases was there any 

substantial loss of information at the family level, 

and that in two cases the sample groupings related 

more closely to the gradient of pollution at the 

phylum level than at the species level. Warwick 

tentatively suggested that “antbropogenic effects 

modify community composition at a higher taxonomic level than natural environmental variables, 

which influence the fauna more by species replacement.” Warwick’s paper appears to be the 

only published work to support the use of higher 

taxonomic groups for analysis purposes. In cases 

where only major taxa level data have been 

collected (e.g., m and Tetra Tech, 1988), it has 

been difficult to determine differences in community structure between impacted areas and reference areas, and to establish causes of community 

alterations. Although it would be a cost-saving 

approach, use of higher taxonomic levels to assess 

benthic communities is currently not an accepted 

approach in the United States. 

92.2 ApplicabiIity of Method to Human 

Health, Aquatic Life, or Wildlife 

Protection 

The assessment of benthic community structure is directly applicable to the protection of 

aquatic life. Because benthic organisms are 

aquatic, assessments of benthic community structure provide a direct measure of the condition of 

aquatic life. Furthermore, because bentbic organisms are consumed by other aquatic organisms 

(e.g., fish), assessing the condition of benthic 

communities provides information on other aquatic 

organisms. 

Assessment of benthic community structure is 

both directly applicable to the protection of some 

wildlife (e.g., wading shorebirds that feed on the 

benthic infauna) and indirectly applicable to the 

protection of other wildlife (e.g., fisheating 

wildlife). A substantial decrease in abundance of 

benthic organisms may result in the loss of food 

and a reduction in the value of certain habitat to 

wildlife. For example, distrtbutions of demersal 

fishes have been shown to be affected by changes 

in the composition of benthic infaunal communities (e.g., see Kleppel et al., 1980), as has the 

distribution of the starfish Astiopecten vedi 

(Striplin, 1987). 

Assessment of benthic community structure 

may be directly or indirectly applied to the protection of human health. When changes in community structure are caused by the presence of toxic 

contaminants, the bioaccumulation of those contaminants in more tolerant species may sometimes 

be postulated. Those contaminated benthic infauna may directly affect human health if they are 

ingested (e.g., shellfish contamination), or may 

indirectly affect human health if contaminants are 

transferred through the food web to humans (e.g., 

consumption of contaminated demersal ffih). 

9.2.3 Ability of Method to Generate 

Numerical Criteria for Specific 

Chcmic8ls 

Benthic community structure as a standalone assessment method cannot presently 

generate numerical criteria for specific chemicals, nor is it likely that it will without extensive 

research. However, it is an integral component 

of other methods that generate numerical criteria 

(e.g., Apparent Effects Threshold, Sediment 

Quality Triad). ?he great number of factors 

influencing benthic community structure at a 

given site generally precludes isolation of chemical-specific effects. 

9-IO 
9-4hine Benthic Community Structure Assessment 

93 USEFULNESS 

Assessment of benthic community structure 

has become a valued tool for determining sediment quality. It is recognized as the only in sitar 

measure that provides information on changes in 

ecological relationships among species that inhabit 

potentially contaminated sediment. Its usefulness 

will continue both as an assessment method on its 

own and as a component of other sediment quality 

assessment tools. 

93.1 Environmental Applicability 

This method is applicable in a variety of 

environments. As a tool for assessing sediment 

quality, it has been used to asess the effects of 

known or suspected contaminants (e.g., industrial 

or municipal discharges, oil spills). The results of 

such studies reveal the geographic extent of the 

problem area and the type and severity of contamination. 

9.3.1.1 S&ability for Different Sediment Tjpes 

Benthic community structure is well suited for 

assessing spatial and temporal effects of chemical 

contamination and/or organic enrichment in a 

variety of sediment types. However, to the extent 

possible, benthic communities occupying different 

types of sediment should not be compared. Considerable research has shown that the structure of 

benthic communities in coarse sediments differs 

from that in fine sediments (see Rhoads and 

Young, 1970, Rhoads and Boyer, 1982). Briefly, 

species recruiting into soft, silty sediments must 

be able to lolerate the deposition of fine particulate material. These environments tend to be 

inhabited by subsurface deposit-feeding organisms, 

whereas sandy environments tend to be inhabited 

by both surface suspension-feeding species and 

subsurface-dwelling species. Therefore, the 

experimental design of a benthic survey must 

reflect that the functiona attributes of benthic 

communities in silty and sandy environments 

fundamentally differ. 

When reference stations are used as the basis 

for determining differences in community stxudure 

between nonimpaded and potentially impacted 

stations, the reference and test stations should 

exhibit, to the extent possible, similar sediment 

characteristics (as well as similar water depths 

because benthic communities naturally vary by 

depth). However, it is not always possible for the 

reference and test stations to have sediment that 

has a similar composition; for example, dredged 

material at a dump site may have different charactefistics than native sediment surrounding the 

dump site. If the experimental design is based on 

sampling the same stations through time to assess 

temporal change, then presumably sediment grain 

size would remain constant. If the objective is to 

sample along a potential gradient of chemical 

contamination or organic enrichment, then all 

stations should have similar grain sizes and water 

depths. However, this is not always possl%le 

because the source of contamination may alter the 

natural grain size dist&ution of the sediments. 

Benthic community structure is also a suitable 

technique for assessing the presence of anaerobic 

sediments caused by poor flushing or excessive 

organic loading. The success of this approach will 

once again hinge on comparing bentbic community structure between stations with similar grain 

sizes and water depths. 

9.3.1.2 Suitability for Digerent Chemical or 

Classes of Chemicals 

Analysis of benthic community structure is 

frequently used to determine effects of chemicals 

present in the sediment. However, it is not used 

as a method to quantify the relative concentrations 

of individual chemicals or classes of chemicals 

present in sediment. Although individual species 

may react to certain chemicals, these reactions are 

not quantifiable at the community level. The 

Apparent Effects Threshold approach (Chapter 10) 

incorporates changes in abundance of major taxa 

for specific chemicals. 

Benthic communities respond predidably to 

general categories of contamination. For example, 

metals contamination of sediments results in 

decreased species diversity (Rygg, 1985a, 198!%, 

9-11 
Sediment Chss$cation Methods Compendium 

1986). Organic enrichment, which leads to an 

increase. in the food supply, generally results in 

increased diversity and abundance at slight to 

moderate levels of enrichment (Pearson and 

Rosenberg, 1978; Rygg, 1986). However, beyond 

some level of organic enrichment, diversity and 

abundance decrease with continued organic loading (Pearson and Rosenberg, 1978). In an area 

receiving both organic enrichment and toxic 

contaminants, it may be difficult to distinguish the 

effects of these forms of pollution from each 

other. Additional research is greatly needed to 

help separate the effects of multiple sources of 

contaminants. 

9.3.1.3 Suitability for Predicting Effects on 

Diflerent 0rganism.s 

Changes in benthic communities that result 

from the presence of organic enrichment or chemical pollutants may be useful indicators of the 

potential effects of that pollution on predators of 

the infauna (see Kleppel, 1982; Striplin, 1987). 

However, using benthic community structure to 

predict specific effects on potential predators (such 

as benthic-feeding fish or shorebirds) may be 

difficult. Information on trophic relationships, 

competition, and predation is often not available. 

The capability to predict the effects of altered prey 

communities on predators may improve with 

research on these topics. Factors such as food 

quality, distribution of food, interactions among 

species, and distribution of prey will all be important components of this research. 

9.3.1.4 Suitability for In-Place Pollutant 

Control 

Benthic community structure has not been 

used to set sediment quality goals or aiteria for 

polluted marine sediments. Benthic communities 

naturally express sufficient spatial and temporal 

variability to eliminate them from consideration as 

a goal or criterion-setting variable. However, 

benthic communities are an integral part of other 

approaches to assess sediment quality (see Chapters 10, and 11, and 12) in which benthic community stkcture k the only in situ biological 

sure. 

9.3.1.5 Suitabili~ for Source Control 

meaBenthic community assessments can provide 

valuable information for certain aspects of source 

control. Benthic communities can assist the 

identification of outfalls that discharge toxic 

chemicals or high organic loads. Depending on 

the nature of tie material being discharged, benthic communilies may be diverse and abundant if 

the material is organically enriched or may be 

depauperate if the material has high levels of toxic 

contaminants. Because benthic communities are 

not currently useful for identifying specific chemicals or classes of chemicals present in the sediment, they are of limited value for identifying 

specific sources of contaminants. 

Following the control of sources, benthic 

community structure may be used to monitor longterm recovery of the receiving environment (Tetra 

Tech, 1988). It is not recommended as an indicator of the immediate effects of controlling sources 

because the sediment will remain contaminated 

until the sediment is actively remediated, or until 

bioturbation and natural deposition of uncontaminated particulates dilute the contaminated 

sediment. Furthermore, this assessmentechnique 

would be useful only in areas where other uncontrolled sources would not obscure sediment recove-ry due to the controlled source. Where source 

control has occurred, or is planned on a regional 

level, establishment of one or more stations for the 

analysis of long-term trends in benthic community 

structure is recommended as a method of monitoring regional sediment recovery. The concentration 

and type of the contaminants and the hydrodynamics of the study area will govern the length of 

time over which recovery will occur (Pereg K, 

1 May 1989, personal communication). 

9.3.1.6 Suitability for Disposal Applications 

Regulations concerning biological testing of 

sediment that is dredged under sections 401 and 

9-12 
9-h4arine Ben&c Communily Stnccture Assessment 

404 of the Clean Water Act do not include assessments of benthic community structure. Benthic 

communities inhabit only the upper layers of 

sediment that will be dredged. Because sediment 

quality near the sediment surface may not reflect 

sediment quality throughout the depth of sediment 

to be dredged, benthic communities are unable to 

provide information that is suitable for assessing 

the entire volume of sediment that will be 

dredged. memica analyses, laboratory bioassays, 

and bioaccumulation studies can, however, be 

used to assess sediment quality throughout the 

dredging depth. Section 102 of the Marine Protection Research and Sanctuary Act does call for 

monitoring of benthic community structure in 

areas where dredged material is disposed. 

The International Joint Commission (IJC) 

recommends use of benthic communities to determine whether areas of concern exist in sediments 

that require dredging (IJC, 1988a, 1988b). However, they do not discuss whether benthic community structure would be used to determine the suitability of dredged material for open-water disposal. 

Analysis of benthic community structure is 

appropriate for postdisposal monitoring of cunfined and unconfined disposal sites and for monitoring recovery of areas that were dredged. As 

part of the Puget Sound Dredged Disposal Analysis (PSDDA) postdisposal monitoring program, 

benthic community structure is used to monitor 

the potential transport of disposed material away 

from the disposal site (SAIC, 1991; Striplin et al., 

1991). The purpose of this aspect of the monitoring program is to determine whether benthic 

communities are altered near the disposal site and, 

if so, whether the changes are due to offsite 

migration of the disposed material. Benthic 

community structure was also incorporated into 

the proposed monitoring program for confined 

aquatic disposal sites to confirm recolonization of 

the clean sediment cap and to monitor cap integrity at the Commencement Bay Near&ore/ 

Tideflats Superfund site in Tacoma, Washington 

(Tetra Tech, 1988). As described earlier, Swartz 

et al. (1980) documented recovery in Yaquina 

Bay, Oregon, following dredging. Rhoads et al. 

(1978) suggested that periodic disturbance such as 

dredging and disposal may enhance benthic 

productivity. 

93.2 General Advantages and Limitations 

General advantages of using benthic community structure to determine sediment quality 

include its inberent capability to provide aa 

ecological basis for evaluation of sediment quality. 

It is an empirical rather than a theorekal approach. However, as with most assessm ent techniques involving field studies, the evaluation of 

benthic communities is costly and time-consuming. The information gained is often not suitable 

for specific management decisions because of the 

lack of numerical management criteria and the 

method’s inability to identify specific chemicals 

responsible for an impact. However, the technique has been incorporated into other predictive 

techniques (see Chapters 10, 11, and 12) that 

provide information more easily used by resource 

managers. 

9.3.2.1 Ease of Use 

Assessments of benthic community structure 

require field collections, extensive laboratory 

work, and data analysis and interpretation by 

trained benthic ecologists. It is difficult to argue 

that the method is easy to use, especially in 

comparison to other methods that rely on established criteria. However, the use of bent& 

community structure as a sediment quality assessment tool is widely accepted, and trained bent& 

ecologists are available throughout the country. 

By using highly experienced individuals to amduct the field, laboratory, and data analysis work, 

potential problems (such as generating “noisy” 

data that obscure real trends, or arriving at different interpretations using the same data) should not 

occur. 

9.3.2.2 Reldive Cost 

The relative cost of conducting an assessmat 

of benthic communities is less than the cost to 

9-13 
Sediment CIass+cafion Methods Compendium 

develop and implement other sediment quality 

assessment techniques such as the Apparent 

Effects Threshold and equilibrium partitioning 

approaches. However, once sediment quality 

values have been generated, the relative cost of 

conducting a benthic survey is greater than the 

cost of analyzing sediment for contaminant concentrations and comparing those data to the values 

to determine sediment quality. Sediment toxicity 

bioassays are generally less costly than analysis of 

replicated benthic samples. Because the Triad 

approach requires synoptic analyses of sediment 

chemistry, sediment toxicity, and benthic communities, it is more costly to implement than simply 

an analysis of benthic communities. It also 

provides broader information from which to 

determine sediment quality. 

The objectives of benthic community assessment programs strongly influence cost by dictating 

the number of stations and number of replicates 

per station. The cost per replicate is relatively 

high (i.e., S400-S 1 ,OOO), but varies greatly depending on the size of the area sampled, the screen 

size, the level of the taxonomic identifications, 

and the environment sampled. 

9.3.2.3 Tendency to Be Conservative 

Benthic community strudure is a moderately 

conservative measure of sediment quality. Because benthic community structure reflects the 

collective response of all species, responses of 

individual species that are susceptible to degradation in sediment quality may not be obvious at the 

community level because of the lack of response 

in olher species that are more tolerant of environmental degradation. Changes to numerous species 

or dominant species must occur before changes at 

the community level are evident. If assessments 

of sediment quality were made using individual 

species instead of communities, they could be 

either conservative by relying on sensitive species 

or not conservative by relying on tolerant species. 

9.3.2.4 Level of Acceptance 

Benthic community assessments have been 

used as a sediment quality assessment tool for 

several decades in North America, Europe, and 

Australia, as well as in South Africa, China, and 

Japan. The method has gained widespread acceptance because of its inherent capability to assess 

sediment quality at the community level, thereby 

documenting ecological response to sediment 

perturbations. 

Many me&ds may be used to analyze benthic community data, as discussed above. Some 

of these methods have gained far wider acceptance 

than have other, sometimes newer, approaches. 

The most widely accepted types of analyses 

include measures of abundance, numbers of taxa, 

diversity, similarity, community classification, and 

the abundance of sensitive and tolerant species. 

Other analytical methods include the log-normal 

distribution (Gray and Mizra, 1979), the use of 

major taxa instead of species-level data (Warwick, 

19@3), and the Infaunal Trophic Index (Word, 

1978, 1980). Each of these may be appropriate 

for certain types of perturbations, but have yet to 

gain widespread acceptance. 

9.3.2.5 Ability to Be Implemented by 

Laboratories with l)pkal Equipment 

and Handling Facilities 

Many laboratories either have the essential 

equipment for conducting benthic community 

surveys or can readily obtain this equipment. 

However, locating qualified taxonomists to oversee the sorting and to identify the organisms may 

be difficult. Taxonomists require several years of 

training and experience before they are considered 

experts in their respective taxonomic fields. They 

also require access to a reference museum of 

verified organisms to assist in their identifications. 

A thorough taxonomic library containing original 

descriptions of species is also an integral component of taxonomic laboratories. 

93.2.6 Level of Efort Required to Generate 

Rest& 

The level of effort required to conduct a 

benthic community survey is dependent on the 

objectives of the program, which may affect the 

number of stations, number of replicates per sta9-14 
9-Mminc Benthic Community Stnrcture A-t 

tion, taxonomic level of the identifications, and 

data analysis procedures. Regardless of those 

objectives, a field effort is required; the samples 

must be sorted, identified, and enumerated; and 

the resulting data must be analyzed. This process 

typically requires several months, but it is not 

unusual for it to require a full year for a very 

large sampling effort, or for a program in which 

the samples require large sorting or identification 

times. For example, the sorting time for samples 

collected from deep water silt and clay may be 

l-2 h, whereas that for samples from shallow 

sandy sites might be 4-6 h because shallow sandy 

areas typically contain more abiotic material. If 

wood chips are present in the sample, then the 

sorting time can easily exceed 12 h, depending on 

the volume of wood chips. 

9.3.2.7 Degree to Wkh Results Lend 

Themselves to Interpretation 

The interpretation of benthic community data 

requires an expert who is familiar with the natural 

history of the fauna and the statistical techniques 

that are routinely used to analyze the data. Interpretation of the many data points generated by this 

approach may require many weeks before meaningful trends are recognized. ‘Ihe inherent variability of benthic communities has so far prevented the development of specific benthic criteria for 

use in assessing pollutant-related trends in sediment quality. 

9.3.2.8 Degree of Environmental Applicability 

The assessment of benthic community structure is a direct measure of the environmental 

effects of pollutants and, as such, is highly applicable as a method to assess sediment quality. Its 

applicability lies in its ability to provide inforrnation on the effects of pollutants on ecological 

processes within the sedimentary environment. 

9.3.2.9 Degree of Accuracy and Precision 

Provided that sufficient funding is available to 

collect and process the necessary numbers of 

replicate samples, analysis of benthic community 

structure is accurate {defined as how well the data 

represent true field conditions) and precise (defied as the consistency and reliability of the 

samples). The resulting data are obtained directly. 

from the populations under study. Other sediment 

quality assessment methods descrtbed in this 

compendium are not direct measures of field 

conditions and therefore are less likely to be as 

accurate and precise. 

Many factors in the design of a benthic community survey directly influence the degree of 

accuracy and precision of the resulting data. 

These factors include station placement, number of 

replicates, appropriateness of reference areas, 

sampler, sieve mesh size, sampling interval, 

quality of taxonomy, and the type and quality of 

the data analysis. The best way to ensure high 

degrees of accuracy and precision is to conduct a 

pilot study in the ares of interest prior to designing a major field survey. The pilot survey will 

provide information on variability within benthic 

communities, which then directly affects the 

required number of replicates and station placement. lhe analysis of data from a pilot study 

may also help generate different hypotheses that 

may alter the sampling and analysis plans to better 

define the communities. 

9.4 S’l’AlWS 

Many methods to assess sediment quality rely 

on benthic community structure as a measure of 

potential ecological effects of pollutants. Benthic 

community structure has been incorporated into 

programs with vastly different objectives because 

the resident biota are sensitive indicators of many 

kinds of environmental perturbations. Aspects of 

the status of benthic community structure as a 

sediment quality assessment tool are discussed in 

this section. 

9.d.l Extent of Use 

Assessment of benthic community structure 

has been a valued tool in marine, estuarine, and 

9-15 
Sediment CZassz+x~tiun Methods Compendium 

freshwater environments for several decades. 

Many of the early programs examined benthic 

communities from an academic viewpoint. Since 

the 197Os, benthic community structure has been 

used as a measure of sediment quality. Since then 

this method has been used to determine the effects 

of municipal effluents, industrial discharges, 

eutrophication, organic enrichment, oil spills, and 

mine tailings disposal (see Section 9.1.1). It has 

also been used to determine the suitability of 

sediments for dredged material disposal, to monitor dredged material disposal sites, and to monitor 

recovery of impacted areas following the cessation 

of contaminant loading. 

9.4.2 Extent to Wbicb Approach Has Been 

Field-Validated 

Because benthic community structure is an in 

situ sediment quality assessment tool, it does not 

require additional field validation. 

9.43 Reasons for Llmlted Use 

Although conducting studies of bentbic community structure is a common practice, the cost 

and amount of time required to generate usable 

results may prevent the method from being implemented by all who could benefit from its use. In 

fact, the method has been deleted from some 

programs due solely to cost (Bilyard, 1987). In 

some situations, costs and time have been reduced 

by taking the identifications only to tbe major 

taxonomic level. This reduction of taxonomic 

detail frequently reduces the usefulness of the 

information (Warwick, 1988), wbicb exacerbates 

a perception by some resource managers that the 

data are too variable to be useful. Detecting 

trends within benthic data is not a simple process. 

However, the proper design and implementation of 

a field survey will radically increase the probability of producing valuable data and results. 

9.4.4 Outlook for Future Use and Amount of 

Development Yet Needed 

The outlook for the future use of benthic 

community structure as a sediment quality assessment tool is particularly bright because of the 

continuing development of new data analysis 

methods by researchers in North America and 

Europe. The objective of these methods is generally to reduce cost or variability within the data by 

relating aspects of the distributions of organisms 

or organism biomass to specific kinds of environmental perturbations. Gray and Mirza (1979) 

determined that the lognormal distribution of 

tidividuals was altered in a predictable manner in 

the presence of slight organic pollution. A more 

recent method for detecting pollution effects on 

marine bentbic communities is the species abundance/biomass comparison (ABC) method developed by Warwick (1986). This method proposes 

that the relationship between the number of 

individuals among species and the distribution of 

biomass among species changes in a predictable 

manner in the presence of organic pollution. 

Beukema (1988) evaluated the ABC method in an 

intertidal habitat in the Dutch Wadden Sea and 

determined that the method “cannot be applied to 

tidal flat communities without reference to longterm and spatial series of control samples.” Yet 

another benthic community assessment method 

that remains under development is the Infaunal 

Trophic Index proposed by Word (1978, 1980). 

That method is based on changes in the feeding 

ecology of benthic infauna in relation to organic 

enrichment. ‘Ibe Benthic Resource Assessment 

Technique, developed by Lunz and Kendall 

(1982), quantifies the effects of changes in benthic 

communities on fish resources. Although the 

BRAT technique is not a direct assessment of 

benthic community stTUcture, it provides important 

information on the relationships among benthic 

communities and higher level predators, and 

describes how those relationships may &ange in 

the presence of pollutants. 

A radically different approach to interpreting 

long-term changes in benthic community structure 

involves use of a se&men1 profile camera. Rhoads and German0 (1986) developed the REMOT!% (remote ecological mapping of the seafloor) system. ‘I%ey use a vessel-deployed sediment-profile camera to photograph vertical sections of the sediment, Although REMOTSS 

9-16 
9”Aurinc Benthic Community Structure Assessment 

cannot determine the species composition of the 

benthic community, it can document relationships 

between organisms and sediment. Rhoads and 

Getmano (1986) characterized the successional 

stages of benthic communities and suggested that 

mapping these stages will permit the detection of 

changes in benthic communities. When this 

information is collected as part of a preliminary 

survey, it can be used to assist in the design of a 

cost-efficient benthic community survey for 

obtaining geochemical and biological information. 

Additional research is needed on some fundamental aspects of benthic community assessment. 

These include the development of guidelines for 

the identification of reference sites or reference 

values and additional studies into the usefulness of 

identifying infauna to various taxonomic levels. 

U.S. EPA is presently examining some aspects of 

these questions through the Clean Water Act 

section 301(h) program, including examination of 

the degree of variability in benthic communities in 

contaminated and reference areas, development of 

a quantitative definition of “balanced indigenous 

populations,” and assessment of the effects of 

overlapping contaminant sources on benthic 

infaunal communities. 

The sediment profile camera has been used for 

a variety of other purposes including assessing the 

relationships between sediment quality and eutrophication (Day ef al., 1987; Revelas ef al., 1987; 

Rhoads, D.C., 1 May 1989, personal communication), monitoring the perimeter of dredged material disposal sites (Rhoads, D.C., 1 May 1989, 

personal communication; Diaz, RJ., 1 May 1989, 

personal communication), and evaluating the 

overwintering habitat of blue crabs in Chesapeake 

Bay (Schaffner and Diaz, 1988). With further 

research, the sediment profile camera may be used 

for other applications concerning aspects of 

benthic community structure and sediment quality. 

9.5 REFERENCES 

Amjad, S., and J.S. Gray. 1983. Use of the 

nematode/copepod ratio as an index of organic 

pollution. Mar. Poll. Bull. 14:178-181. 

Austin, M.P., and P. Grieg-Smith. 1%8. The 

application of quantitative methods to vegetation survey. IL Some methodological problems of data from rain forest. J. Ecol. S&827- 

844. 

Beukema, JJ. 1988. An evaluation of the ABC 

method (abundance-biomass comparison) as 

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9-19 
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9-21 
CHAPTER 10 

Sediment Quality Triad Approach 

peter M. Chapman 

E.V.S. Consultants Ltd. 

195 Pemberton Avenue, North Vancouver, BC, Canada V7P 2R4 

Phone (604) 986-4331, FAX (604) 662-8548 

The Sediment Quality Triad (Triad) approach 

is an effects-based approach to describe sediment 

quality. It typically incorporates measures of 

sediment chemistry, sedimentoxicity, and benthic 

infauna communities, although other variables can 

be used. This combination method is both descriptive and numeric. It is most commonly used 

to describe sediment qualitatively, but has also 

been used to generate chemical-specific sediment 

quality criteria (Chapman, 1986, 1989; Long, 

1989). One application of the Triad approach, the 

Apparent Effects Threshold (AET), is described in 

detail in the following chapter (Chapter 11). 

10.1 SPECIFIC APPLICATIONS 

10.1.1 Current Use 

The Triad approach can be used to determine 

the extent of pollution-induced degradation of 

sediments in a non-numerical, multiple-chemical 

mode (e.g. Chapman et al., 1986, 1987a, 1991a; 

Chapman and Power, 1990; Chapman, 1990). It 

can also be used to determine numerical sediment 

quality criteria directly (e.g. Chapman, 1986, 

1989) and, through manipulations, to determine 

AET values (see Chapter 11). 

The AET is only one possible method of 

evaluating triad data and is directed solely at 

determining numeric sediment quality values 

(Chapman et al., 1991b, 1991c). The triad approach has been used in marine coastal waters on 

the west coast of North America (e.g., Puget 

Sound, San Francisco Bay, and Vancouver Harbor, 

Canada), in the Gulf of Mexico, in freshwater 

environments including the Great Lakes, and in 

the North Sea (Long and Chapman, 1985; chapman, in press; Chapman et al., 1986, 1987a, in 

press; Chapman and Power, 1990, Cross et al., 

1991, in review). Current uses of the Triad 

approach are summarized in Table 10-1 and 

discussed in Section 10.3.1, Environmental 

Applicability. 

10.1.2 Potential use 

The Sediment Quality Triad approach can also 

be used to meet the following objectives: 

• To identify problem areas of sediment 

contamination where pollution-induced 

degradation is occurring; 

• To prioritize and rank degraded areas and 

their environmental significance; and 

• To predict where such degradation will 

occur based on levels of contamination 

and toxicity. 

The Triad approach can be used in any number of situations and is not restricted to aquatic 

sediments. For example, it can be used in water 

column work with phytoplankton and in terrestrial 

hazardous waste dump studies with other organisms of concern. Other uses are described in 

Section 10.3.1. A complete description of the 

Triad in the context of integrated assessments i

provided in Chapman et al., 1991b. 

10.2 DESCRIPTION 

10.2.1 Description of Method 

The Triad approach consists of three comportents (Figure 10-1): 

• Sediment chemistry--to measure chemical 

contamination; 
Sediment Classification Methods Compendium 

Table 10-1. Current Uses of the Sediment Quality Triad Approach. 

PS = Puget Sound, various locations (Long and Chapman, 1985). 

GM= Gulf of Mexico, oil platform (Chapman et al., 1991a; Chapman and Power, 1990). 

SF = San Francisco Bay, various locations (Chapman etal., l986, l987a). 

VH = Vancouver Harbor. Canada, various locations (Chapman etal., 1989; Cross et al, 1991; Cross et al; in 

review). 

FW = Various freshwaternvironments (Malueg et al., 1984; Chapman unpublished data; Rogers. North Texas 

State, unpublished data; Wiederholm et al., 1987). 

NS = North Sea (Chapman, In press; Chapman et al, in press). 

• Sediment bioassays-to measure toxicity; 

• In situ biological variables - to measure 

in situ alteration (e.g., a change in benthic 

community structure). 

The three components provide complementary 

data. No single component of the Triad approach 

can be used to predict the measurements of the 

other components. For instance, sediment chemistry provides information on contamination but not 

on biological effects. Sediment bioassays provide 

direct evidence of sediment toxicity. However, 

the laboratory conditions under which bioassays 

are conducted may not accurately reflect field 

conditions of exposure to toxic chemicals. In situ 

alteration of resident biota measured by infauna 

community analyses provides direct evidence of 

contaminant-related effects in the environment, but 

only if confounding effects not related to pollution 

(e.g., competition, predation, recruitment cycles, 

sedimentype, salinity, temperature, recent dredging) can be excluded. In particular, because the 

toxicity of a chemical substance in sediments may 

vary with its concentration and with the conditions 

within a specific sediment, the importance of any 

particular concentration of a chemical or suite of 

chemicals in sediments cannot be determined 

solely from chemical measurements. Sediment 

conditions include grain size, organic content, pH, 

Eh, chemical form, and presence of other 

chemicals. 

The three components of the Triad approach 

integrate chemical and biological response data. 

They also provide the strongest evidence for 

identifying pollution-induced degradation. For 

instance, if there are high levels of sediment 

contamination, toxicity, and biological alteration, 

the burden of evidence indicates degradation. 

Conversely, low levels of sediment contamination, 

toxicity, and biological alteration indicate nondegraded conditions. Conclusions that can be 

drawn from intermediate responses are listed in 

Table 10-2. 

10-2 
I&Triad Approach 

SEDIMENT CHEMISTRY 

Figure 10-l. Conceptual Model of the Sediment Quality Triad. 

The Triad combines data from chemistry, toxicity bioassays, and in situ studies. Chemistry and 

bioassay estimates are based on laboratory measurements with field-wkcted sediments. In situ 

studies generally include, but are not limited to, measures of benthic community structure. Areas 

where the three facets of the Triad show the greatest overlap (in terms of either positive 01 

negative results) provide the strongest ‘data for determining sediment quality criteria. 

10.2.1.1 Objectives and Assrunprions 

The objectives of tie Triad approach are to 

independently measure sediment contamination, 

sediment toxicity, and biological alteration, and 

then use the burden of evidence to assess sediment 

quality based on all three sets of measurements. 

The following assumptions apply: 

n The approach allows for (1) the interactions between contaminants in complex 

sediment mixtures (e.g., additivity, antagonism, synergism); (2) the adions of 

unidentified toxic chemicals; and (3) the 

effect of environmental faders that influence biological responses (including toxicant concentrations). 

m Selected chemical contaminant concentrations are appropriate indicators of overall 

chemical contamination. 

H Bioassay test results and values of selected benthic community structure variables 

are appropriate indicators of biological 

effects. 

Tbe.se components are presently often treated 

in an additive manner, with each having equal 

103 
Sediment Clnssfktion Methuds Compendium 

Table l&2. Possible Conclusions Provided by Using the Sediment Ouality Triad Approach. 

due to toxic chemicals 

‘+ = Measured difference between test and amtrol or refersnm amditbm. 

- I No measurable differencs between test and control or reference oonditkns. 

weight because there is insufficient information 

available to assign weightings. 

10.2.1.2 Level of Eflort 

Ideally, the Triad approach would be based on 

the use of synoptic data. Sediments for analysis 

of toxicity should come from the same cornposited 

homogenate, as originally detailed by Chapman 

(1988), ideally from field ralher than solely laboratory test replicates. Benthic infauna samples 

should be collected- at the same sampling locations. Chemistry and bioassay sediments are 

collected (usually by remote grab), transferred to 

a solvent-rinsed glass or stainless steel bowl, and 

thoroughly homogenized by stirring with a glass 

or stainless steel spatula until textural and color 

homogeneity are achieved. The homogenized 

sediments are then placed in appropriate sampling 

container% In generat, chemistry and bioassay 

samples should include field rather than laboratory 

replication. Benthic infaunal samples are collected 

at the same location. In the absence of initial 

sampling to determine the optimum level of 

replication at a site, five field replicate benthic 

samples are recommended per station (see aapter 

8, Methods). Coincident rather than synoptic 

sampling is possrble (e.g., Long and Chapman, 

1985); however, spatial heterogeneity ln sediment 

contamination and toxicity make such data diffrcult to interpret (Swark d al., 1982). 

Adequate quality QA/QC measures must be 

followed in all aspects of the study, from field 

sampling through laboratory analyses and data 

entry. Detailed QA/QC procedures are available 

through international (e.g., Keith cf al., 1983) and 

regional publications (e.g., Terra Tech, 1986a). 

The first component of the Triad invoIves 

identification and quantification of inorganic and 

organic contaminants present in the sediments. 

aiemical analytes measured are generally restricted by quipmeat, technology, and the availability of funds and facilities. IALA concerns and 

existing data also affect target analytes measured. 

Cost, if a factor, must be balanced against the 

need for an analytical database sufficiently large 

104 
lo-Triad Approach 

Table 10-3. Example Analytes and Detection Units for Use in the 

Chemistq Component of Sediment Quality Triad Approach. 

Arsenic 

Iron 

Chromium 

Wwr 

Cadmium 

Lead 

Mercury 

Nickel 

Silver 

Selenium 

Zinc 

LPAH’ 5 

Benz0 (a) pyrene 10 

Benzo(e)pyene 10 

Benz0 (a) anthracene 10 

Chrysene 10 

Dibenzoanthracene 16 

Fluoranthene 5 

Pyrene 5 

The detection limits are the Instrumental estimates. Actual detection llmtts may be higher bcKauseofm8trbcescts. 

‘TOC = total organic carbon. 

bAVS = AVS methodology is described by the U.S. EPA (1991); modiications are expected. Contact 

Christopher Zarba at (202) 475-7326 to obtain latest protocok 

’ LPAH = low-molecular-weight polycylic aromatic hydrocarbons (Includes acemaphthene, anthracene, naphthaha 

and methylated naphthalenes, fluorene, phenanthrene, and methylated phenanthrenes). 

‘PCBs= polychlorinated biphenyls. 

‘PCP = pentachlorophenol. 

‘TCP = tetrachlorophenol. 

to allow determination of the presence (or ab- Total organic carbon and grain size are measence) of known toxicants of concern. sured to provide a basis for normalizing the data 

An example of some of the types and classes to different type-s of sediments. Acid volatile 

of compounds required to provide a reasonable sulfides (AVS) provide information for de&m& 

characterization of chemica1 contamination is ing metals availability from sediments. &proshown in Table 10-3. stanol, an indicator of human waste, can be 
Sediment Classi&ztiun M&u& Compendium 

Table l&4. Possible Static Sediment Bioassayr. 

Rhepoxyniw abronk& 

(adutt amphipod) 

Bivalve Larvae 

development 

NeaHhes sp. 

Survival, avoidmce 

survival, developtnti 

Hydella azteca 

(adult amphipod) 

Daphnia magna 

(water tlea) 

Si~rvival, avoidance 

Survival, reproduction 

Estuarinc Weten 

Eohaustorius estuarius 

(aduit amphipod) 

Survival, avoidance 

’ Note: Other options indude but are not necessarily restricted to Ampeka sbdifa, Co@~iun w/u&tw, Grandidienella iaponica, Foxipheikrs xixim@us 

measured to differentiate sewage inputs from 

industrial inputs. 

The second Triad component involves identification and quantification of toxicity based on 

laboratory tests using field-collected sediments. 

Ideally, one would test the toxicity of the sediments to all ecologically and commercially important fauna living in or associated with the 

sediments. For logistical reasons, a small number 

of bioassays is conducted to cover as wide a range 

as possible of organism type, life cycle, exposure 

route, and feeding type. The number of tests 

undertaken is affected by the same constraints as 

those mentioned for sediment chemistry analyses. 

Possible static sediment bioassays that provide 

a reasonable characterization of the degree of 

toxicity are shown in Table 10-4. Obvious omissions from this list include full life-cycle chronic 

tests, and genotoxic or cytotoxic response tests. 

Such tests merit consideration for inclusion when 

proven accepted methods become available (e.g., 

Ung and Buchman, 1989). 

The final Triad component involves the 

evaluation of in sihr biological alteration. Generally, this component is provided by benthic infauna community data because benthic organisms 

are relatively sessile and location-specific. Hist+ 

pathology of bottom fish has also been used for 

this Triad component (Chapman, 1986), but for 

areawide rather than site-specific studies, because 

these fish are relatively mobile. Several variables 

in combination are effective in characterizing 

benthic community structure for the Triad 

approach: numbers of taxa, numerical dominance, 

IO-6 
lo-Triad Appmch 

total abundance, and percentage composition of 

major taxonomic groups. In the marine environment, this last category includes any or all of 

polychaetes, amphipods, molluscs, and echinoderms. In the freshwater environment, oligocbaetes, chironomids, and other major insect 

groups would fit into the last category. 

Sediment chemistry, toxicity, and benthic 

infauna data are combined in the Triad approach 

to asses the degree of degradation of each station 

and of each site (see Figure 10-l). All data are 

compared on a quantitative basis and can be 

normalized to reference site values by converting 

them to ratio-to-reference (RTR) values as described by Chapman et al. (1986, 1987a) and 

Chapman (1990). The reference site chosen 

(either u priori or a posteriori) is generally the 

least contaminated site of those sampled, and 

ideally its sediment and other characteristics (e.g., 

water depth) would be similar to those of the 

other sites. To determine RTR values, the values 

of specific variables (e.g., normalized concentration of a particular metal, percent mortality in a 

particular bioassay, number of taxa) are divided by 

the corresponding reference values. This process 

normalizes the data so that they can be compared 

even when, for instance, there are large differences 

in the units of measurement. The reference site 

may be a single station (whose RTR value is 1.0 

by definition) or an area containing several stations for which data are averaged. 

The RTR criterion is based but does not 

depend on the assumption that the reference site 

concentrations are indicative of reference or 

background conditions. The degree to which 

chemical concentrations are elevated above the 

mean reference concentrations at a selected site is 

used as the criterion for selecting chemicals most 

likely to be anthropogenically enriched and of 

concern. An index of contamination can be 

calculated for each station by separately determining RTR values for groups of similar chemicals 

(e.g., metals, PAH, chlorinated organ&) and then, 

assuming additivity, combining these values as a 

single mean chemistry RTR value. Similarly, an 

index of toxicity can be calculated by combining 

bioassay RTR values as a single mean value. 

Finally, an index of biological alteration can be 

calculated in the same manner as is toxicity, using 

benthic community structure data. The indexes of 

contamination can be used to rank stations. These 

summary ranks can also be compared with the 

ranks generated using the sediment bioassay and 

infaunal data. 

The composite RTR values for each Triad 

component can also provide useful visual indexes. 

These values can be plotted on scales with a 

common origin and placed at 120 degrees from 

each other such that each of the three values 

becomes the vertex of a triangle. The relative 

degree of degradation is derived by calculating 

and comparing the areas of the triangles for each 

station or site. Examples of such triaxial plots are 

shown in Figure 10-2, for the eight possible situations shown in Table 10-2. These plots also 

provide a visual guide to the characteristics of 

background or reference stations. Because reference data usually involve a site containing more 

than one reference station, RTR comparisons 

should also be made against individual reference 

stations. Alden (1992) provides a method for 

determining confidence limits for such triaxial 

plots. Non-RTR methods of Triad data anabsis 

are outlined in Section 10.2.1.2.3, Types of Data 

Required. 

10.2.1.2.1 Type of Sampling Required 

As described, synoptic sampling is preferred 

for all three Triad components. Any reasonable 

sampling procedure can be used if it provides 

suitable sediment samples for quantifying sediment contamination, toxicity, and biological 

alteration. To date, studies have used remote 

samplers such as a 0.1-m’ Van Veen grab operated 

from a vessel. 

10.2.1.2.2 Methods 

Typical variables included in the chemical 

analyses and sediment bioassays are listed in 

Tables 10-3 and 10-4, respectively. Details for 

benthic infauna analyses are provided in aapter 

8. Although unit costs vary, costs are generally 

on the order of $1,500 for three separate replicated 

(n=5) sediment bioassays, Sl,!KKl for unreplicated 

lo-7 
Sahent Clakfication Methods Compendium 

TOXlCrry 1 l I 

ToxmrY 1+r 1 *I CONTAMWATDN 

Figure l&2 Sedlnwnt Qurtity Trial Ttiudal plots for tha Eight Possible Skmtions Shown In Trek 10-2 

730 SedimW Quality Tiiad determined, in the axampie situation, for each of the e@ht pa&&k 

outcomes described in T&/e 10-2. Toxicity, contamination, and aiteratiw~ are shown nomWzsd 

to Ratio-tcMeferences values as described by Chapman et al. (1986, 1987a), 1.0 = rem 

conditions. Note that the exact symmetry in these examples w&d not be mutitwly expcted In 

actual studies. 
lo-Triad Approach 

chemical analyses, and $2,500 for replicated (n=5) 

benthos. 

10.2.1.2.3 Types of Data Required 

Standard measurements of chemistry, toxicity, 

and biological alteration are required. These 

measurements can then be combined, as described 

above. Detailed data calculations and analyses are 

as folIows: 

Data Calculations - Benthic Data 

n Calculate/determine endpoints 

l taxa richness 

l total abundance 

l numerical dominance 

l species diversity 

l mean abundances of all species of major taxa (e.g. polychaetes, amphipods, 

chironomids, oligochaetes) 

n Cluster Analysis 

l e.g., using mean numbers of individuals 

per taxa present at each station tested. 

Data Calcuiatbns - Chemistry 

n Bulk concentration normalized to dry 

weight 

n Organic carbon normalized concentration 

of organic compounds 

n Normalize to percent fines, sand, silt, and 

clay 

w AVS normalized concentration of metals 

(DiToro et al., 1990; Dewitt et al., 1990) 

m Summarize means, standard deviations, 

ranges for each parameter at each site. 

Data Calculations - Bioassay 

I Between station differences in mean response, ANOVA, multiple comparison 

kStS. 

l Paired comparison with control response. 

n Comparison of mean response with lower 

prediction limit (LPL) (Dewitt et al., 

1988); this comparison addresses possible 

grain-size effects on amphipods. 

Non-RTR Methoa!s of Triad Data Analysis 

The traditional reduction technique of calculating RTRs (by translating m&ant measures to 

proportions of comparable values obtained for the 

reference site) has the following problems (Cross 

et al., 1991; Cross et al., in review): 

n Substantial loss of information during the 

conversion of multivariate data into single 

proportional indexes; 

n Loss of any spatial relational information; 

n Inability to statically assess significance 

of spatial impacts; and 

l Requirement of an appropriate reference 

station. 

Xn addition, Triad results could be strongly 

influenced by the presence of unmeasured toxic 

contaminants that may or may not covary with 

measured chemicals (Chapman, 1990). The RTR 

approach is useful in specific situations and with 

defined limitations; however, the following options are useful for reducing or removing the 

problems identified. 

Ranking-In addition to RTRs, rankings can also 

be assigned to biological, chemical, and toxicological data for statistical comparisons of the data. 

Using the chemistry data as an example, the 

sample with the lowest level of a chemical is 

scored as 1 and the highest is scored with a 

number that is equal to the number of time periods or samples that are to be ranked. Tied data 

should be scored by calculating an average of the 

tied ranks. Each site will have a rank for each 

biological, chemical, and toxicological parameter. 

An overall mean rank for each site can be cahlo-9 
St&men t Classijicn tion Methods Compendium 

lated using each of the parameters. This effectively determines how each site compares to each of 

the other sites. 

Average ranks for biological, chemical, and 

toxicological data can also be calculated and can 

be compared using Kendall’s coefficient of concordance (Zar, 1984). High concordance will 

indicate that biological, chemical, and toxicological parameters are changing in the same direction 

(improving or degrading). Low concordance will 

indicate that biological, chemical, toxicological 

data are changing independently of each other. 

Multivariate Analysis-Multivariate analysis 

comprises data matrix preparation, analysis independent of the Triad components, analysis concurrent with the Triad components, and Mantel’s test. 

Each of these is briefly described here. 

Data Matrix Preparation 

For each Triad component, data are standardized to common units where possible and incorporated into separate matrices for analysis and 

interpretation. 

Ben thos: Data are abundance of each taxon 

per grab sample; transformed to log 

(x+1). 

Chemistry: Values less than the detection limit 

are omitted to maintain the integrity 

of the matrix. Remaining data are 

log-transformed. 

Bioassay: Because of the number of independent bioassays and differing endpoints (e.g., mortality, avoidance, 

reburial, etc.), these data cannot be 

standardized to common internal 

units. Various transformations 

(arsine square root, log, etc.) may 

be used as required. 

Independent AM~J.JS~S of the Triad Components 

Each matrix is analyzed separately to determine environmental impact as provided by each 

independent approach. Community classification 

analysis may be performed for each data matrix 

using cluster analysis. “Boot-strapping” techniques developed by Nemec and Brinkhurst 

(1988a, 1988b) can be used to test whether clusters of samples differ significantly from each 

other. 

Concurrent A~lysis of the Triad Corqwments 

The ecological ordination technique, principal 

components analysis (PCA), can be used to examine relationships between benthos community 

structure, toxicology, and the physical-chemical 

attributes of the bottom sediments, (Cross et al., 

1991, in review). PCA is used to reduce the 

multidimensionality of the benthos data, creating 

two variables (principal component or PC) from 

the original matrix of many variables (taxon 

abundances). These PCS can then be correlated 

with PQ derived from physicalchemical data a 

bioassay results, or with individual physical or 

chemical parameters. High correlations among 

PCs from the three Triad components indic@e 

agreement or concordance of impact assessments. 

Correlations of PCs from benthic data (or 

bioassay data) with individual chemical parameters 

can be used to assess or develop sediment quaIity 

aiteria. The impacts associated with existing 

aiteria can be expressed as a PC score for benthic 

data, calculated from a regression of these scores 

on chemical concentrations. Sediment quality 

criteria could also be developed by predicting the 

chemical concentration associated with a significant impact on the benthic community, provided 

that “significant impact” could be unequivocally 

associated with a particular PC score or range of 

scores. 

Mantel’s Test 

Another method that can be used to determine 

whether different components of the Triad are 

related is Mantel’s test (Mantel, 1967; Legentire 

and Fortin, 1989). Mantel’s test uses a randomization procedure that akuiates the probability 

that two distance matrices are more similar than 

would be expected by chance alone. Multivariate 

lo-10 
lO-Triad Approach 

(or univariatc) distance between each of the sites 

(observations) can be calculated using data from 

each component of the Triad. For example, to 

develop a distance matrix based on toxicity test 

results, each of the toxicology variables would be 

used to develop the distance. Similar matrices 

would be calculated for benthos and chemistry 

data. 

The randomization procedure ensures that the 

relationships between two distance matrices are 

real and not spurious. The distance between two 

stations (A and 8) is always partially related to 

the distance between these two and other stations 

(e.g., A and C, B and C). Mantel’s test avoids the 

possibility of spurious correlations by calculating 

correlations between the two matrices based on 

random samples, and comparing the actual correlation with the distribution based on the random 

samples. 

10.2.1.2.4 Necessary Hardware and Skills 

Appropriate sampling equipment and trained 

field and laboratory personnel are required for 

chemical analyses, toxicity testing, and benthic 

infaunal analyses. Although the equipment required can be both costly and sophisticated, it is 

commonly necessary for sediment contamination 

investigations. The necessary equipment, facilities, and expertise are generally available through 

a wide variety of government, university, cornmercial, and private facilities. 

10.2.1.3 Adequacy of Documenfdon 

Documentation for use of this method is 

provided by Long and Chapman (1985), Chapman 

(1986, 1989, 1990), and Chapman el al. (1986, 

1987a, 1991a, 1991b). Other investigators have 

also successfully applied this method (cf. Chapman et al., 1991c). 

10.2.2 Applicability of Method to Human 

Health, Aquatic Life, or Wildlife 

Protection 

This approach is directly applicable to the 

protection of aquatic life. To date, only benthic 

invertebrates and fish have been used to assess in 

si.2~ biological effeds and sediment toxicity. 

Protection of aquatic life may indirectly proted 

wildlife (e.g., wading birds feeding on benthos) 

and humans (e.g., via consumption of aquatic life). 

The approach can be directly applicable to human 

health and wildIife protection if the Triad comPonents are redirected towards issues such as 

bacteria1 contamination and toxic contaminant bioaccumulation. For instance, Triad could be used 

in three ways to address bacterial problems: 

(1) measure bacterial contamination in water or 

sediment, (2) measure bacterial diseases or concentrations in tissues, and (3) perform laboratory 

tests to quantify relationships between sediment/water concentrations and effects. Toxic 

contaminant bioaccumulation could be addressed 

by these uses of the Triad approach: (1) measure 

toxic contaminant concentrations in water or 

sediment, (2) measure bioconcentration/biomagnification in lissues, and (3) perform laboratory 

tests to determine effects related to bioconcentration and biomagnification. 

10.23 Ability of Method to Generate 

Numerical Criteria for Specific 

Chemicals 

The Triad approach has been used to generate 

criteria for three contaminants: lead, PAH, and 

PCBs (Chapman, 1986). These criteria were 

developed in Puget Sound by examining large data 

sets to identify contaminant areas and concentrations that were associated with no or minima1 

biological effects. The criteria fall within a factor 

of 2 to 10 of values generated for the-se contaminants by the screening-level concentration (see 

Chapter 11, Section ll.l.l.), the AET approach 

(see Chapter ll), and laboratory toxicity methods 

(Chapman el ol., 1987b). As detailed by Chapman (1989), the AET application of the Triad 

concept provides criteria for benthic infauna and 

each bioassay conducted, whereas the latter cornbines all bioassay and in sihc biological effects 

data to provide a single value, interpretation, or 

analysis. However, there has been little work 

since Chapman (1986) on development of the 
Sediment Class$cation Methods Compendium 

Triad approach for the production of numerical 

sediment quality criteria separate from AET. 

10.3 USEFULNESS 

103.1 Environmental Applicability 

Although the Triad approach is both laborintensive and expensive, its strengths render it 

extremely cost-effective for the level of information provided. First, it provides empirical 

evidence of sediment quality (based on observation, not theory). Second, it allows ecological 

interpretation of physical, chemical, and biological 

properties (i.e., interpretation of how these relate 

to the real environment). Third, it uses a preponderance-ofevidence approach rather than relying 

on single measurements (i.e., all the data are 

considered). Because of the comprehensive nature 

of Triad studies, additional follow-up studies are 

usually not necessary. Finally, the data generated 

by the Triad approach can be used to generate 

effects-based classification indexes. 

The Triad approach enables investigators to 

estimate the size of degraded and nondegraded 

areas. It also provides a test of the quality of 

reference areas (i.e., do contamination or biological effects occur?). Standards in the form of 

sediment quality criteria (Chapman, 1986, 1989; 

ITI, 1988a, 1988b) can be set from the contaminant concentrations that are always associated with 

effects, using the AET application of the Triad. 

The Triad approach also provides the information 

necessary to describe the ecological relationships 

between sediment properties and biota at risk from 

sediment contamination. 

The Triad approach has been used in dredging 

studies to support dredged material disposal siting 

and disposal decisions (aapman, unpublished). 

In multiplying the relative degree of degradation 

at a site by the volume of sediment to be dredged, 

investigators can compare different sites, provided 

that the same reference area is used. This comparison helps investigators determine whether 

dredging will affect useful habitat or result in 

material unacceptable for ocean disposal. Similarly, potential disposal sites can be compared with 

each other and with the material to be dredged, 

and then compared to acceptability criteria for 

various uses and options. This application of the 

Triad approach replaces similar but less useful 

comparisons based solely on the total mass of 

chemical contaminants to be dredged. 

In areas where benthic communities have been 

eliminated or drastically changed because of a 

natural event (e.g., storms, oxygen depletion) or 

physical anthropogenic impact (e.g., recent dredging, boat scour), the other two Triad components 

(sediment chemistry and toxicity) provide information when conventional univariate approaches 

would prove deficient. Such cases emphasize the 

need to use knowledge of an area in making any 

type of environmental assessment, including the 

Sediment Quality Triad. 

The Triad approach can be used to discern 

and ultimately to monitor regional trends in 

sediment quality. Such information is necessary 

to delineate areas that are excessively contaminated with toxic chemicals affecting the biota and, 

therefore, most in need of remedial action. Pilot 

studies of this nature have been conducted in 

Puget Sound and San Francisco Bay (Long and 

Chapman, 1985; Chapman, 1986; Chapman cl al., 

1986, 1987a) and in Europe (e.g., Chapman, in 

press; Chapman et (II.; in press). 

3.1.1 Suitability for Ditrerent Sediment 

VP= 

The Triad approach can be used with all 

sediment lypes, including sands, muds, aerobic 

sediments, and anaerobic sediments. It includes 

sediment characterization with physical parameters 

[e.g., ,gtain size, acid volatile sulfides (AVS), and 

total organic carbon (TOC)] that may be important 

in interpreting the Triad compounds. For example, caution must be used in interpreting the 

results of toxicity tests in sediments that remain 

anaerobic in the laboratory despite aeration. 

Specifically, organisms will die from lack of 

oxygen, making it difficult to distinguish that 

mortality loom toxicity due to high concentrations 

of contaminants. 

10-12 
10.3.1.2 Suirability for Different Chemicals or 

Classes of Chemicals 

The Triad approach can be used with all 

chemicals or classes of chemicals, provided that 

bioassay organisms and tests are appropriate for 

all chemicals. For this reason, a battery of bioassay tests 1s recommended. Caution must be used 

when testing sediment extracts that may be specific to certain chemical classes. Interpretation of 

the results must be restricted to only those 

chemicals. 

10.3.1.3 Suitability for Predicting Eflects on 

Diflerent Organisms 

Application of the Triad approach can be 

limited by the organisms in the environment if the 

in situ effects are determined primarily by the 

same species used in the bioassay tests. In other 

words, all biological effects data are based on a 

single species. ln such cases, independence of the 

infaunal community analyses and bioassay test 

results cannot be assumed. Hence, more than one 

bioassay test is recommended. Ideally, the tests 

would include a wide variety of organisms, life 

stages, feeding types, and exposure routes. 

10.3.1.4 Suitability for in-Place Pollutant Control 

The Triad approach provides a comprehensive 

approach to in-place pollutant control because it 

allows for assessment of all potential interactions 

between chemical mixtures and the environment. 

This method is comprehensive because it includes 

the measurements of multiple chemicals as well as 

the potential toxic effects of both measured and 

unmeasured chemicals. 

10.3.1.5 Suitability for Source Control 

The Triad approach is as suitable for sourcontrol as it is for in-place pollutant control. It 

can be an environmental complement to toxicity 

reduction evaluation (TRE) programs that involve 

chemical and toxicity investigations of sediments, 

and effluents and other discharges. 

10.3.1.6 Suitability for Disposal Applicatians 

The Triad approach has been used for disposal 

applications, including Navy Homeporting work in 

San Francisco Bay. in that study, the Triad 

approach clearly separated potential dredge sites 

from one another in terms of the relative level of 

pollution. Although the Triad was not used in the 

final decision because of other considerations, 

decision-makers were able to use information 

provided by the Triad to compare the suitability of 

dredging and disposal options. 

103.2 General Advantages and Limitations 

The following are the major advantages of the 

Triad approach: 

Combines three separate components to 

provide a preponderance-of-evidence 

approach; 

Does not require a priori assumptions 

concerning the specific mechanisms of 

interaction between organisms and toxic 

contaminants; 

Can be used to develop sediment quality 

values (including criteria) for any measured contaminant or a combination of 

contaminants, including both acute and 

chronic effects; 

Provides empirical evidence of sediment 

quality; 

Can be used for any sediment type; 

Allows ef3logical interpretation of both 

physical-chemical and biological properties; and 

Does not usually required follow-up when 

a complete study is conducted. 

The following are the major limitations to the 

Triad approach: 

lo-13 
Sediment Classification Methods Compendium 

n Statistical criteria have not been fully 

developed for use with the Triad approach 

(but see Section 10.2.1.2.3, Types of Data 

Required); 

n Rigorous criteria for calculating single 

indexes from each of the sediment chemistry, bioassay, and in situ biological 

effects data sets have not been developed 

(but may not be required); 

n A large database is required; 

n If the approach is used to determine single-chemical criteria, results could be 

strongly influenced by the presence of 

unmeasured toxic contaminants that may 

or may not covary with measured 

chemicals; 

n Methods for sediment bioassay testing 

need to be standardized; 

l Sample collection, analysis, and interpretation are labor-intensive and expensive; and 

n The choice of a reference site is often 

made without adequate information on 

how degraded the site may be. 

10.3.2.1 Ease of Use 

The Triad approach is relatively easy to use 

and understand. The concept is straightforward. 

A high level of chemical and biological expertise 

is required to obtain the data for the three separate 

Triad componenls. However, many laboratories or 

groups of laboratories possess the required 

expertise. 

10.3.2.2 Relative Cost 

Relative cost can be evaluated in either dollars 

or environmental damage. The Triad approach 

may not prevent environmental damage, but it can 

be used to identify contaminated areas for future 

remediation. In terms of dollars, the Triad approach requires substantial resources to be implemented properly, although step-wise, tiered use of 

Triad components is possible. Measured against 

the potential environmental damage due to toxic 

contamination and the costs of remediation, the 

Triad approach can be extremely cost-effedive. 

10.3.2.3 Tendency to Be Conservative 

The Triad approach provides objedive data 

with which to determine and sometimes to predict 

environmental damage. Its predictive ability 

allows for, but does not require, conservatism on 

the part of the decision-makers. 

10.3.2.4 Level of Acceptance 

The Triad approach is gaining a high level of 

acceptance in various parts of North America and 

in Europe (Forstner et al., 1987; Chapman, in 

press). In addition, Canada has conducted Triad 

studies in Vancouver to determine the suitability 

of this approach for implementation of the new 

Canadian Environmental Protedion Ad (Cross et 

al., 1991; Cross et al., in review). 

10.3.2.5 AbiIity to Be Implemented by 

L&oratories with Typical Equipment 

and Handling Facilities 

All aspects of the Triad approach (i.e., benthic 

infaunal studies, sediment chemistry analyses, 

sediment toxicity bioassays) can be conducted by 

any competent, specialist laboratory that is reasonably well equipped. The major requirements are 

adequate QNQC procedures for chemical measurements; appropriate detection limits; and, for 

biological analyses, taxonomic experts and a 

taxonomic reference library or museum. 

10.3.2.6 Level of Efort Required to Generate 

Results 

Different levels of effort will generate different levels of results. For instance, results can be 

generated by simply measuring one or two chcmicals, determining the number of infauna present, 

and conduding a single sediment toxicity bioas10-14 
IO-Triad Approach 

say. However, the applicability of these results 

may be severely limited. Consequently, multiple 

chemicals including inorganic and organic compounds should be measured, and in situ biological 

alteration and sediment toxicity should be measured multiple times. Although it is possible to 

use previously collected nonsynoptic data to 

derive results in a “paper” study (Long and fiapman, 1985), fieldwork dnd synoptic sampling 

generate the most useful results. 

10.3.2.7 Degree to Which Results Lend 

Themselves to Interpretation 

Beyond the general conclusions noted in Table 

10-2, expert judgment is required to implement 

and interpret the Triad approach. In particular, the 

definition of “minimal” and “severe” biological 

effects is required to establish chemical-specific 

criteria. The Triad approach reflects the complexity of the issues that must be addressed to assess 

environmental quality. 

10.3.2.8 Degree of Environmental Applicability 

As discussed, the Triad approach has an 

extremely high degree of environmental applimbility (see Section 10.3.1). 

10.3.2.9 Degree of Accuracy and Precision 

The accuracy and precision of the Triad 

approach have not been quantitatively determined. 

It is expected to have a high degree of accuracy 

and precision, although these parameters will vary 

with those of the constituent components. 

10.4 STATUS 

10.4.1 Extent of Use 

,Development of the formalized Triad concept 

has occurred relatively recently (ktg and Chapman, 1985; Utapman, 1986, 1990; Chapman et 

al., 1986, 1987a, 1988, 1991a). The Triad approach has been used directly to establish sediment quality criteria (Chapman, 1986) and, 

through data manipulations, to determine AFT 

values for sediment quality uiteria (Tetra Tech, 

1986a; PTI, 1988a, 1988b). 

The Triad has been used to identify spatial 

and temporal trends of pollution-induced degradation. indexes developed using the Triad approach 

can be numeric (as described in Chapter 11 for the 

AFT application of the Triad concept) or primarily 

descriptive (see Figure 2, Chapman ef al., 1987a). 

In either case, the Triad approach provides an 

objective identification of sites where contamination is causing discernible harm (cf. Power et 

al., 1991). 

10.4.2 Extent to Which Approach Has Been 

Field-Validated 

Because the Triad approach measures in situ 

biological alteration in the field, field validation is 

an integral part of each complete Triad 

investigation. 

10.43 Reasons for Limited Use 

As previously described, the Triad approach is 

being used in the United States, Canada, and 

Europe for marine, estuarine, and freshwater areas. 

It is not being used in small projects because of 

the cost and expertise required for full 

implementation. 

10.4.4 Outlook for Future Use and Amount 

of Development Yet Needed 

The following areas of the Triad approach 

require development: 

n Determining tire appropriateness of the 

various endpoints of different bioassays, 

sekted chemical contaminants, selected 

measures of benthic community structure, 

and other potential measures of in siru 

biological alteration; 

m Determining the appropriateness of an 

additive treatment of the data (e.g., summing bioassay responses to provide a 

single index for toxicity); 

lo-15 
Sdimcnt Cfass~tidn lAdha& Gqxndium 

B Further development of statisticai criteria; 

n Development of rigorous criteria for determining, where and if appropriate, composite indexes for each of the three Triad 

components; and 

n Continued standardization of methods for 

sediment toxicity bioassays. 

Even without development of these areas, the 

Triad approach provides valuable information. 

The argument has been made (Chapman er al., 

1986, 1987a) that the Triad approach provides 

objective information on which to judge the extent 

of pollution-induced degradation. For this reason 

the Triad approach will likely be used much more 

widely in future. 

10.5 REFERENCES 

Alden, R. W. II. 1992. Uncertainty and sediment 

quality assessments: I. Confidence limits for 

the Triad. Environ. Toxicol. Chem. 11:637- 

644. 

Urapman, P.M. 1986. Sediment quality criteria 

from the Sediment Quality Triad - an example. Environ. Toxicol. Chem. 5: 957964. 

Chapman, P.M., R.N. Dexter, S.F. Cross, and 

D.G. Mitchell. 1986. A field trial of the 

Sediment Quality Triad in San Francisco Bay. 

NOAA Technical Memorandum NOS OMA 

25. National Oceanic and Atmospheric Administration, San Francisco, CA. 127 pp. 

Chapman, P.M., R.N. Dexter, and E.R. Long. 

1987a. Synoptic measures of sediment contamination, toxicity and infatmal community 

structure (the Sediment Quality Triad) in San 

Francisco Bay. Mar. Ecol. Prog. Ser. 37:75- 

%. 

Chapman, P.M., R.C. Barrick, J.M. Neff, and RC 

Swartz. 1987b. Four independent approaches 

to developing sediment quality criteria yield 

similar values for model contaminants. Environ. Toxicol. Chem. 6~723-725. 

Chapman, P.M. 1988. Marine sediment toxicity 

tests. pp. 391402. In: Chemical and Biological Characterization of Sludges, Sediments, Dredge Spoils, and Drilling Muds. J.J. 

Lichtenberg, F.A. Winter, C.I. Weber, and L 

Fradkin (eds.). ASTM STP 976. American 

Society for Testing and Materials, Philadelphia, PA 

Chapman, P.M. 1989. Current approaches to 

developing sediment quality criteria. Environ. 

Toxicol. Chem. 8: 589-599. 

Chapman, P.M., C.A. McPherson, and K-R. Munkittrick. 1989. An asessment of the Ocean 

dumping tiered testing approach using the 

Sediment Quality Triad. Unpublished report 

prepared for Environmental Protection Canada. E.V.S. Consultants, North Vancouver, 

BC., cirnada. 

Chapman, P.M., and E.A. Power. 1990. Sediment toxicity evaluation. American Petroleum 

Institute Publication No. 4501. 209 pp. 

Chapman, P.M. 1990. The Sediment Quality 

Triad approach to determining pollutioninduced degradation. Sci. Total Environ. 

9718:815-825. 

Chapman, P. M. In press. Pollution status of 

North Sea sediments-An international scientific study. Mar. Ecol. Prog. Ser. 

Chapman, P.M., RN. Dexter, H.k Andersen, and 

B.A. Power. 1991a. Evaluation of effects 

associated with an oil platform, using the 

Sediment Quality Triad. Environ. Toxicol. 

C&em. l&407-424. 

Chapman, P. M., E. A. Power, and G. A Burton, 

Jr. 1991b. pp. 313-340. Chapter 14: Integtative assessments in aquatic ecosystems. In: 

Contaminated Sediment Toxicity Assessment. 

G. A Burton Jr. (ed.). &is Publishers, 

Qaelseq Michigan. 

Chapman, P.M., E.R: Long, R C Swarzz, T.H. 

Dewitt, and R. Pastorok. 1991~. Sediment 

toxicity tests, sediment chemistry and benthic 

ecology & provide new insights into the 

significance and management of contaminated 

sediments - a reply to Robert Spies. Environ. 

Toxicol. Chem. 10:1-4. 

Chapman, P.M., RC Swartz, B. Roddie, H 

Phelp, P. van den Hurk and R Butler. In 

press. An international wmperison of ttedi- 
IO-Triad Approach 

ment toxicity tests in the North Sea. Mar. 

Ewl. Prog. Ser. 

Cross, S.F., J.M. Boyd, P.M. Chapman, and R.O. 

Brinkhurst. 1991. A multivariate approach for 

defining spatial impacts using the Sediment 

Quality Triad. p. 886. In: Proceedings of the 

17th Annual Aquatic Toxicity Workshop, 

P.M. Chapman, F. S. Bishay, E. A. Power, K 

Hall, L. Hardking. D. McLeavy, M. Nassichuk 

and W. Knapp (eds.). Can. Tech. Rept. Fish. 

Aquat. Sci. 1774. 

Cross, S. F., J. M. Boyd, P. M. Chapman, and R. 

0. Brinkhurst. (In review). A multivariate 

approach to assessing the spatial extent of 

benthic impacts established using the Se&- 

ment Quality Triad. Environ. Toxicol. Chem. 

Dewitt, T. H., G. R. Distwortb, and R. C. 

Swartz. 1988. Effects of natural sediment 

features on survival of the Phoxocephalid 

amphipod, Rhepoxynius abronius. Mar. 

Environ. Res. 24:99-124. 

Dewitt, D. M., J. D. Mahony, D. J. Hansen, K. J. 

Scott, M. B. Hicks, S. M. Mayr, and M. S. 

Redmond, 1990. Toxicity of cadmium in 

sediments: the role of acid volatile sulfide. 

Environ. Toxiwl. Chem. 9:1487-1502. 

DiToro, D.M., J.D. Mahony, DJ. Hansen, KJ. 

Scott, M.B. Hicks, S.M. Mayr, and M.S. 

Redmond. 1990. Toxicity of cadmium in sediments: The role of acid volatile sulfide. Environ. Toxiwl. Chem. 9: 1487-1502. 

Forstner, V.U., F. Ackerrnann, J. Alberti, W. 

Calmano, F.H. Frimmel, K.N. Komatzki, R. 

Leschber, H. Rossknecht, U. Schleichert, and 

L. Tent. 1987. Qualitatskriterien fur Gewassensedimente - Allgemeine Problematik und 

internationaler stand der Diskussion. WasserAbwasser-Forsch 2054-59. 

Keith, L.H., W. Crummett, J. Deegan, Jr., RA. 

Libby, J.K. Taylor, and G. Wentler. 1983. 

Principles of environmental analysis. AnaI. 

Chem. 552210-2218. 

Legendre, P and MJ. Fortin. 1989. Spatial pattern 

and ecological analysis. Vegetatio 80:107- 

138. 

Long, E. R. 1989. The use of the Sediment Quality Triad in classification of sediment wntamination. pp. 78-93. In: Marine Board, National 

Research Council Symposium/Workshop on 

contaminated marine sediments. 

Long, E.R., and M.F. Bucbman. 1989. An evaluation of candidate measures of biological 

eSfeds for the National Status and Trends 

ROgKilTl. NOAA Technical Memorandum 

105 pp. NOS OMA 45: National Oceanic and 

Atmospheric Administration, Rockmille, MD. 

Long, E.R., and P.M. Chapman. 1985. A sediment quality triad: measures of sediment 

contamination, toxicity and infaunal colfLII1unity composition in Puget Sound. Mar. Poll. 

Bull. 16~tO5-41’5. 

Malueg, K.W., G.S. Schuytema, D.F. Krawczyk, 

and J.H. Gakstatler. 1984. Laboratory sediment toxicity tests, sediment chemistry and 

distributions of benthic macroinvertebrates in 

sediments from the Keweenaw Waterway, 

Michigan. Environ. Toxiwl. aem. 3231 

242. 

Mantel, N. 1967. The detection of disease clusteting and generalized regression approach. 

Cancer Res. 27:200-209. 

Nemec, A.F.L, and R-0. Brinkhurst. 1988a. 

Using the bootstrap to assess statistical significance in the cluster analysis of species abundance data. Can. J. Fish. Aquat. Sci. 45:%5- 

970. 

Nemec, A.F.L, and R.O. Brinkhurst. 1988b. The 

Fowlkes-Mallows statistic and the comparison 

of two independently determined dendrograms. Can J. Fish. Aquat. Sci. 45:971-975. 

Power, E. A., K. R. Munkittrick, and P. M. Chap 

man. 1991. An ecological impad assessment 

framework for decision making related to 

sediment quality. pp. 48-64. In: Aquatic 

Toxicity and Risk Assessment: Fourteenth 

Volume. M. A. Mayers and M. G. Barron 

(eds.). ASTM STP 1124. American Society 

for Testing and Material, Philadelphia, PA. 

FI’I Environmental Sentices, Inc. 1988a. Sediment quality values refinement: Tasks 3 and 

5 -1988 update and evaluation of Puget Sound 

AET. Unpublished report prepared for Tetra 

Tech, Inc. for the Puget Sound Estuary Program, EPA Contract No. 68-02-43441. PTI 

Environmental Services, Inc., Bellevue, WA. 

lo-17 
Sediment Classij~tion Methods Compendium 

PTI Environmental Services, Inc. 1988b. Briefing 

report to the EPA Science Advisory Board: 

the Apparent Effects Threshold approach. 

Unpublished report prepared for Battelle 

Columbus Division, EPA Contract No. 68-03- 

3534. PTI Environmental Services, Inc., 

Bellevue, WA. 

Swartz, R.C., WA. DeBen, KA. Sercu, and J.O. 

Lunberson. 1982. Sediment toxicity and the 

distribution of ampbipods in Commencement 

Bay, Washington, USA. Mar. Poll. Bull. 

13:359-364. 

Tetra Tech. 1986a. Recommended protocols for 

measuring selected environmental variables in 

Puget Sound. Prepared for the Puget Sound 

Estuary Program, U.S. Environmental Protection Agency, Region X, Seattle, Washington. 

Tetra Tech, Inc., Bellevue, WA 

Tetra Tech. 1986b. Development of sediment 

quality values for Puget Sound. Prepared for 

Resource Planning Associates and U.S. Army 

Corps of Engineers, Seattle District, for the 

Puget Sound Dredged Disposal Analysis 

Program. Tetra Tech, Inc., Bellevue, WA. 

U.S. EPA. 1991. Analytical method of determination of acid volatile sulfide in sediment. U.S. 

Environmental Rote&on Agency, aiteria 

and Standards, Washington, DC. 

Wiederholm, T., A-M. Wiederholm, and G. Milbrink 1987. Bulk sediment bioassays with 

five species of fresh-water oligocltaetes. 

Water, Air and Soil Pollut. 36: 131-154. 

Zar, J. H. 1984. Biostatistical Analysis, 2d ed. 

Prentice-Hall, Englewood aiffs, NJ. 

10-18 
CHAPTER 11 

Apparent Effects Threshold Approach 

John Malek 

0ffice of Puget Sound, U.S. Environmental Protection Agency Region X 

1200 Sixth Avenue, Seattle, WA 98101 

(208) 553- 1286 

In the Apparent Effects Threshold (AET) 

approach, empirical data are used to identify 

concentrations of specific chemicals above which 

specific biological effects would always be expected. Following the development of ART values for 

a particular geographic area, they can be used to 

predict whether statistically significant biological 

effects are expected at a station with known 

concentrations of toxic chemicals. 

11.1 SPECIFIC APPLICATIONS 

11.1.1 Current Use 

At present, the AET approach is being used 

by several programs as guidelines for the protection of aquatic life in Puget Sound. These guidelines are the culmination of cooperative planning 

and scientific investigations that were initiated by 

several federal and state agencies in the early and 

mid-1980s. 

Three programs and applications of the AET 

approach are highlighted below. Notably, all 

these programs involve an element of direct 

biological testing in conjunction with the use of 

AET values, in recognition of the fact that no 

approach to chemical sediment quality values is 

100 percent reliable in predicting adverse biological effects- An underlying strategy in many of 

these programs was to develop two sets of sediment quality values based primarily on AET 

values: 

• One set of values identifies low chemical 

concentrations below which biological 

effects are improbable. 

• A second set of values identifies higher 

chemical concentrations above which 

multiple biological effects are expected. 

The programs incorporate direct biological testing 

in concentration ranges between these two extremes to serve as a “safety net” (i.e., to account 

for the uncertainty of chemical predictions) for 

potential adverse effects or anomalous situations 

at “moderate” chemical concentrations. 

Commencement Bay Nearshore/Tideflats 

Superfund Investigation 

Commencement Bay is a heavily industrialized harbor in Tacoma, WA. Recent surveys have 

indicated over 281 industrial activities in the 

nearshore/tideflats area. Comprehensive shoreline 

surveys have identified more than 400 point and 

nonpoint source discharges in the study area, 

consisting primarily of seeps, storm drains, and 

open channels. A remedial investigation (RI) 

under Superfund, started in 1983, revealed 25 

major sources contributing to sediment contamination, including major chemical manufacturing, 

pulp mills, shipbuilding and repair, and smelter 

operations. Adverse biological effects were found 

in sediments adjacent to these sources. 

The AET approach was developed during the 

course of the RI to assess sediment quality using 

chemical and biological effects data [i.e., depressions in the number of individual benthic taxa, 

presence of tumors and other abnormalities in 

bottom fish, and several laboratory toxicity tests 

(amphipod mortality, oyster larvae abnormality, 

bacterial bioluminescence)]. AET values were 

also used in the subsequent feasibility study (FS) 

to identify cleanup goals and define volumes of 

contaminated sediment for remediation. The AET 

values used in the FS were generated from a 

reduced set of biological effects indicators, which 

comprised depressions in total benthic abundance, 

amphipod mortality, oyster larvae abnormality, 

and bacterial luminescence.. 
Sediment Classification Methods Compendium 

Puget Sound Dredged Disposal Analysis Program 

In 1985, the Puget Sound Dredged Disposal 

Analysis (PSDDA) program was initiated to 

develop environmentally safe and publicly acceptable options for unconfined, open-water disposal 

of dredged material. PSDDA is a cooperative 

program conducted under the direction of the U.S. 

Army Corps of Engineers (Corps) Seattle District, 

US. EPA Region X, the Washington Department 

of Ecology (Ecology), and the Washington Department of Natural Resources (WDNR). AET 

values were used to develop chemical-specific 

guidelines to determine whether biological testing 

on contaminated dredged material is needed. 

Results of the biological testing help determine 

suitable disposal alternatives. 

Above a specified chemical concentration (i.e., 

the screening-level concentration or SLC) biological testing is required to determine the suitability 

of dredged material for unconfined, open-water 

disposal. Based primarily on AET values for 

multiple biological indicators, a higher “maximum 

level concentration” was also identified. Above 

this latter concentration, failure of biological tests 

is considered to be predictable. However, an 

optional series of biological tests can be conducted 

under PSDDA to demonstrate the suitability of 

such contaminated material for unconfined, openwater disposal (Phillips et al., 1988). 

Urban Bay Toxics Action Program 

The Urban Bay Toxics Action Program is a 

multiphase program to control pollution of urban 

bays in Puget Sound. The program includes steps 

to identify areas where contaminated sediments 

are associated with adverse biological effects, 

specify potential pollution sources, develop an 

action plan for source control, and form an action 

team for plan implementation. Initiated in 1984 

by Ecology and U.S. EPA Region X’s Office of 

Puget Sound, the program is a major component 

of the Puget Sound Estuary Program (PSEP). 

Substantial participation has also been provided by 

the Puget Sound Water Quality Authority (Authority) and other state agencies and local governments. Major funding and overall guidance for 

the program is provided by U.S. EPA Office of 

Wetlands, Oceans and Watersheds. 

In the PSEP urban bay program, AET values 

are used in conjunction with site-specific biological tests during the assessment of sediment contamination to define and rank problem areas. 

Source control actions are well under way, but 

sediment remediation has not yet begun at any of 

the sites (PTI, 1988). 

11.1.2 Potential use 

The AET approach to determining sediment 

quality can also be used as follows: 

• To determine the spatial extent and relative priority of areas of contaminated 

sediment; 

• To identify potential problem chemicals in 

impacted sediments and, as a result, to 

focus cleanup activities on potential 

sources of problem contaminants; 

• To define and prioritize laboratory studies 

for determining cause-effect relationships; 

and 

• With appropriate safety factors or other 

modifications, to screen sediments in 

regulatory programs that involve extensive 

biological testing. 

Proposed regulations for sediment contamination 

are currently under review in Puget Sound. These 

regulations may include use of AET values to 

develop statewide sediment quality standards. 

Ecology is currently developing a suite of sediment management standards, as mandated by the 

Puget Sound Water Quality Authority (1988) in its 

1989 Management Plan. The proposed standards 

are based in part on AET values. Development of 

these standards (Becker et al., 1989) relies heavily 

on the past and ongoing efforts described in 

Section 11.1.1 and involves active participation by 

Ecology, U.S. EPA, the Authority, WDNR, the 

Corps (Seattle District), and various public interest 

groups. The draft regulation currently under 

11-2 
II-AET Appwch 

development affects only sediments in Puget 

Sound. As additional data become available from 

other locations, the adopted regulation will eventually be broadened and modified to include the 

entire state. 

11.2 DESCRIPTION 

11.2.1 Description of Method 

AET values are derived using a straightforward algorithm that relates biological and chemical data from field-collected samples. For a given 

data set, the AET for a given chemical is the 

sediment concentration above which a particular 

adverse biological effect (e.g., depressions in the 

total abundance of indigenous benthic infauna) is 

always statistically significant (PsO.05) relative to 

appropriate reference conditions. The calcutalion 

of an AET for each chemical and biological 

indicator is conducted as follows: 

(1) Collect “matched” chemical and biological 

effects data-Conduct chemical and biological effects testing on subsamples of 

the same field sample. (To avoid unaccountable losses of benthic organisms, 

benthic infaunal and chemical analyses 

are conducted on separate samples collected concurrently at the same location.) 

(2) Identify “impacted” and “nonimpacted” 

stations-Statistically test the significance 

of adverse biological effects relative to 

suitable reference conditions for each 

sediment sample. Suitable reference 

conditions are established by sediments 

exhibiting very low or undetectable concentrations of any toxic chemicals, an absence of other adverse effects, and physical characteristics that are directly comparable with those of the test sediments. 

(3) Identify AET using only “nonimpacted” 

stations-For each chemical, the AET can 

be identified for a given biological indicator as the highest detected concentration 

among sediment samples that do not 

exhibit statistically significant effects. (If 

the chemical is undetected in all nonimpacted samples, then no AET can be 

established for that chemical and biological indicator.) 

(4) Check for preliminary m-Verify that 

statistically significant biological effects 

are observed at a chemical concentration 

higher than the AET, otherwise, the AET 

should be regarded only as a preliminary 

minimum estimate. 

(5) Repeat Steps (l)-(4) for each biological 

indicator. 

The AET approach for a group of field-collected sediment samples is shown in Figure 11-l. 

The samples were collected at various locations 

and were analyzed for (1) toxicity in a laboratory 

bioassay and (2) the concentrations of a suite of 

chemicals, including lead and 4-methylphenol. 

Based on the results of bioassays conducted on the 

sediments from each station, two subpopulations 

of all sediments are represented by bars in the 

figure: 

n Sediments that did not exhibit statistically 

significant (P~0.05) toxicity relative to 

reference conditions (“nonimpact& stations) and 

n Sediments that exhibited statistically 

significant (P&05) toxicity in bioassays 

relative to reference conditions (“impacted” stations). 

Over the observed range of concentrations for 

these sediment samples (horizontal axis in Figure 

11-l), the sediments fall into two groups for each 

chemical: 

8 At low to moderate concentrations, significant sediment toxicity occurred in some 

samples, but not in others. 

n At concentrations above an apparent 

threshold value, significant sediment 

toxicity occurre~J in all samples. 
Sediment C&ssfbtion h4ethods Compendium 

Lead 

SP-14 

4 

660 ppm M-16 

iMPACTED 4 

I otzzmmmmmm a 0 G I 

NONIMPACTED AET 

I , I111111, I I I1 ,,,I, I 111,111, 1 I I ItIll, I , ! I,,,,, 

1 10 100 1000 10000 100000 

INCREASING CONCENTRATION ~-& 

OH- 0 

oCH3 4-Methylphenol 

3600 ppb SP-14 

NONIMPACTED A&T 

I I I1111111 I I I III111 I I I 411111 I I I I II111 I I I II1111 I I I Ill111 

1 10 100 9000 WOO0 looooo looooo 

INCREASING CONCENTRATlON - 

Figure 11-l. The AET approach for 8 group d fid~locted wdiment aam* 

The AET approach applied fo sediments tested for lead and I-methylphenol a~~~~?taUons and 

toxicity response during bioassays. 
II-AET Approach 

The AET value is defined for each chemical 

as the highest concentration of that chemical in 

the sediments that did nol exhibit sedimenl toxicity. Above this AET value, significant sediment 

toxicity was always observed in the data set 

examined. Data are treated in this manner to 

reduce the weight given to samples in which 

factors other than the contaminant examined (e.g., 

other contaminants, environmental variables) may 

be responsible for the biological effect. 

For each chemical, additional AET values 

could be defined for other biological indicators 

that were tested (e.g., other bioassay responses or 

depressions in the abundances of certain indigenous benthic infauna). 

11.2.1.1 Objectives and Assutnprions 

The objective of the ART approach is to 

identify concentrations of contaminants that are 

associated exclusively with sediments exhibiting 

statistically significant biological effects relative to 

reference sediments. AET value generation is a 

conceptually simple process and incorporates the 

complexity of biological-chemical interrelationships in the environment without relying on CI 

priori assumptions about the mechanisms of these 

interrelationships. Although the ART approach 

does not require specific assumptions about 

mechanisms of the uptake and toxic action of 

chemicals, it does rely on more general assumptions regarding the interpretation of matched 

biological and chemical data for field-collected 

samples, as described below: 

n For a given chemical, concentrations can 

be as high as the AET value and not be 

associated witb statistically significant 

biological effects (for the indicator on 

which the AET was based). 

n When biological impacts are observed at 

concentrations below an ART value for a 

given chemical, it is assumed that the 

impacts may be related to another chemical, chemical interactive effects, or other 

environmental factors (e.g., sediment 

anoxia). 

n The ART concept is consistent with a 

relationship between increasing concentrations of toxic chemicals and increasing 

biological effects (as observed in laboratory exposure studies). 

The assumptions in interpreting environmental 

data are demonstrated below with actual field data. 

Using Figure 1 l-l as an example, sediment from 

Station SP-14 exhibited severe toxicity, potentially 

related to a greatly elevated concentrations of 4- 

methylphenol (7,400 times reference levels). The 

same sediment from Station SP-14 contained a 

relatively low concentration of lead that was well 

below the AET for lead (Figure 11-l). Despite 

the toxic effects associated with the sample, 

sediments from many other stations with higher 

lead concentrations than Station SP-14 exhibited 

no statistically significant biological effects. 

These results were interpreted to suggest that the 

effects at Station SP-14 were potentially associated with 4-methylphenol (or a substance with a 

similar environmental distribution) but were less 

likely to be associated with lead. A converse 

argument can be made for lead and 4-methylphenol in sediments from Station RS-18. 

Applied in this manner, the ART approach 

helps to identify measured chemicals that are 

potentially associated with observed effects at 

each biologically impacted site and eliminates 

from consideration chemicals that are far less 

likely to be associated with effects (i.e., the latter 

chemicals have been observed at higher concentrations at other sites witbout associated biological 

effects). Based on the results for lead and 

4-methylphenol, bioassay toxicity at five of the 

impacted sites shown in the figure may be associated with elevated concentrations of 4-methylphenol, and toxicity at eight other sites may be 

associated with elevated concentrations of lead (or 

similarly distributed contaminants). 

As illustrated by these results, the occurrence 

of biologically impacted stations at concentrations 

below the AET of a single chemical does not 

imply that ART values in general are not protective against biological effects, only that single 

chemicals may not account for all stations with 

biologica effect& By developing ARTS for 
sediment C&ss@Aon Methods Compendium 

multiple chemicals, a high percentage of all 

stations with biological effects are accounted for 

with the AET approach (see Section 11.3.2.9 and 

USEPA, 1988). 

AETs can be expected to be more predictive 

when developed from a large, diverse database 

with wide ranges of chemical concentrations and 

a wide diversity of measured chemicals. Data sets 

that have large concentration gaps between stations and/or do not cover a wide range of ooncentrations must be scrutinized carefully (e.g., to 

discern whether chemical concentrations in the 

data set exceed reference concentrations) to determine whether AET generation is appropriate. 

11.2.1.2 Level of Eflort 

11.2.1.2.1 Type of SampIing Required 

Collection of field data for initial generation 

of AETs is a labor-intensive and capital-intensive 

process. The exact level of sampling effort 

required depends on the amount and variety of 

data collected (e.g., the number of samples collected, the diversity of biological indicators that are 

tested, and the range of chemicals measured). 

One means of minimizing these costs is to compile existing data that meet appropriate quality 

assurance criteria. There are no definitive requirements for the size and variety of the database, 

although a study of the predictive abilities of the 

AET approach with Puget Sound data (Barrick et 

al., 1988) resulted in the following recommendations for data collection: 

n Collect or compile chemical and biological effects data from 50 stations or more 

(and from suitable reference areas). 

n Bias the positioning of stations to ensure 

sampling of various contaminant sources 

(e.g., urban environments with a range of 

contaminant sources and, preferably, with 

broad geographic distribution) over a 

range of contaminant concentrations (preferably over at least l-2 orders of magnitude). 

n Conduct chemical tests for a wide range 

of chemical classes (e.g., metals, nonionic 

organic compounds, ionizable organic 

compounds). To generate AETs on an 

organic &on-normalized basis, total 

organic carbon (TOC) measurements are 

required in all sediments. 

m Ensure that detection limits of <lOO ppb 

(lower if possible) are attained for organic 

compounds. High detection limits (i.e., 

insensitive analyses) can obscure the 

occurrence of chemicals at low to moderate concentrations; as noted previously, 

only detect@ data are used in AET calculations. Metals are naturally occurring 

substances, and most metals concentrations typically exceed routine de-ion 

limits. 

The only strict requirement for field sampling 

of data for AET generation is the collection of 

“matched” chemical and biological data (as dc 

saibed at the beginning of Section 11.2.1). 

Matched data sets should be used to reduce the 

possibility that uneven (spatially variable) sediment amtamination could result in associating 

biological and chemical data that are based on 

dissimilar sediment samples. Because the toxic 

responses of stationary organisms (e.g., bioassay 

organisms confined to a test sediment, or infaunal 

organisms largely confined to a small area) are 

assumed to be affected by direct association with 

contaminants in the surrounding environment, it is 

considered essential that chemical and biological 

data be collected from nearly identical subsamples 

from a given station. 

1X2.1.2.2 Methods 

Methodological details for the generation of 

AET values are d+bed at the beginning of 

Se&ion 11.2.1. 

11.2.1.23 ‘I)pes of Data Required 

rIJw0 fundamental kinds of data analysis are 

required for AET generation: 

n Statistical analysis of the significance of 

biological effects relative to reference 
II-AET Approach 

conditions (i.e., classification of stations 

as impacted or nonimpacted for each 

biological indicator) and 

n Generation of an AET value for each 

chemical and biological indicator (essentially a process of ranking stations based 

on chemical concentration). 

Additional kinds of data analysis needed for 

AET generation are quality assurance/quality 

control (QA/QC) review of biological and chemical data, and evaluation of the appropriateness of 

reference area stations. These topics have been 

described elsewhere {e.g., Belier et al., 1986; 

Barrick et al., 1988). 

The AET method does not intrinsicaIly require 

a specific method of statistical analysis for determination of significance of biological effects 

relative to reference conditions. Existing Puget 

Sound AETs have relied largely on pairwise 

t-tests; details of statistical analyses performed for 

the generation of Puget Sound AET have been described elsewhere (USEPA, 1988; Banick ef al., 

1988; Beller et al., 1986). For example, the 

following steps were used to determine tbe statistical significance of amphipod mortality bioassay 

results (Swartz et al., 1985) in field-collected 

sediments: 

All replicates from all stations in the 

reference area used for each study were 

pooled, and a mean bioassay response and 

standard deviation were calculated. 

Results from each potentially impacted 

site were then compared statistically with 

the reference conditions using painvise 

analysis. 

The F,, test (Sokal and Rohlf, 1%9) was 

used to test for homogeneity of variances 

between each pair of mean values. 

If variances were homogenous, then a 

t-test was used to compare the two means. 

If variances were not homogenous, then an 

approximate t-test (Sokal and Rohlf, 1969) 

was used to compare the two means. 

n Statistical significance was tested with a 

pairwise error rate of 0.05 to ensure consistency among studies of differing sample sizes. 

Data analyses that have been applied to other 

biological indicators are described elsewhere 

(Beller et al., 1986; Barrick et al., 1988). Notably, comparisons to reference conditions were 

somewhat more complicated for benthic infaunal 

abundances than for sediment bioassays. For 

benthic infaunal comparisons, reference data for 

each potentially impacted site were categorized so 

that comparisons were made with samples collected during the same season, at a similar depth, and 

whenever possible, in sediments with similar 

particle size characteristics (i.e., percentage of 

particles ~64 pm) as those of the potentially impacted site. In this manner, statistical comparisons 

were normalized to account for the influence of 

three of the major natural variables known to 

influence the abundance and distribution of bcnthic macroinvertebrates. All benthic data were 

also log-transformed so that data distributions 

conformed to the assumptions of the parametric 

statistical tests that were applied. Additional data 

treatment methods presented elsewhere (Barrick et 

al., 1988) are not discussed further herein, because 

they are not considered intrinsic to the AET appreach, but rather are options to address poteutially unusual matrices or biological conditions. 

11.2.1.2.4 Necessary Hardware and Skills 

The primary skills required for AET generation are related to the development of the biological/chemical database. Expertise in environmental 

chemistry is required to evaluate chemical data 

quality, and the need for normalization of chemical data and related factors. Biological and 

statistical expertise are required for the determination of statistical significance. For benthic data in 

particular, evaluation of appropriate reference 

conditions and knowledge of benthic taxonomy 

and ecology are necessary. 

Computers are recommended for the efficient 

generation of AET values. A menudriven database (SEDQUAL) has been developed for U.S. 

11-7 
Sediment Class~icution Methods Compendium 

EPA Region X that is capable of a number of data 

manipulation tasks, including the following: 

(1) storing chemical and biological data, (2) calculating AET values, (3) comparing a specified set 

of AET to stored sediment chemistry data to 

identify stations at which adverse biological 

effects are or are not predicted, and (4) based on 

such comparisons, calculating the rate of correct 

prediction of biological impacts. The SEDQUAL 

system, which requires an IBM-AT compatible 

computer with a bard disk, has been documented 

in detail in a user’s manual (Nielsen, 1988). The 

SEDQUAL database currently includes stored data 

from Puget Sound (over 1,000 samples, not all of 

which have biological and chemical data). 

1X.2.1.3 Adequacy of Documentation 

Various aspects of the AET approach have 

been extensively documented in reports prepared 

for U.S. EPA and other regulatory agencies, as 

listed below and in the reference list: 

n 

l 

n 

n 

11.2.2 

Generation of Puget Sound AET values 

and evaluation of tbeir predictive ability 

(Beller et al., 1986; Banick et al., 1988); 

Data used to generate Puget Sound AET 

values (appendices of Beller et al., 1986 

and field surveys cited in Beller et al., 

1986 and Barrick et al., 1988); 

Briefing report to the U.S. EPA Science 

Advisory Board (USEPA, 1988); and 

Policy implications of effects-based marine sediment criteria (PTI, 1987). 

Applicability of Method to Human 

Healtb, Aquatic Life, or Wildlife 

ProtectA 

The AET approach has been designed for use 

in evaluating potential adverse impacts to aquatic 

life associated with chemical contamination of 

sediments. By empirically determining the association between chemical contamination and adverse 

biotogical effects, predictions can be made regarding the levels of contamination that are always 

associated with adverse effects (i.e., the AET 

values). These critical levels of contamination can 

then be used to develop guidelines for protezthg 

aquatic life (e.g., sediment quality values). AETs 

can be developed for any kind of aquatic organism 

for which biological responses to chemical toxicity 

can be measured. The protectiveness of the AET 

can therefore be ensured by evaluating organisms 

and biological responses with different degrees of 

sensitivity to chemical toxicity. For example, 

evaluations of metabolic changes (i.e., usually a 

very sensitive biological response) in a pollutionsensitive species would likely result in AET 

values that are lower and more protective than 

evaluations of mortality (i.e., generally a less 

sensitive response) in a more pollution-tolerant 

species. The protectiveness of AETs am also be 

ensured through the application of “safety factors.” 

For example, to be protective of chronic biological 

responses, a factor based on an acute-chronic ratio 

could be applied to AETs developed on the basis 

of acute biological responses. 

11.23 Ability of Method to Generate 

Numerical Critcrla for Speclfk 

Cbemicds 

The AI3 approach is not intrinsically limited irt 

application to specific chemicals or chemical groups. 

In general, the approach can be used for dremicals 

for which data are available. However, when using 

a specific data set to generate AETs, it is preferable 

that AT3 generation be limited to chemicals with 

wide concentration ranges (e.g., ranging from 

reference concentrations to concentrationsnear direct 

sources) and/or with appropriate detection frequentits (e.g., greater than 10 detections). A partial list 

of chemicals for which AETs have been developed 

is presented in Table 11-l. 

113 USEFULNESS 

113.1 Env&onmcatal AppUcabUity 

113.1.1 S&w for Di$erent Sediment 7)pes 

‘Ilte AET approach can be applied to any 

sediment type in saltwater or freshwater environments for which biological tests can be conducted. 
II-AET Approach 

Table 11-l. Selected Chemicals for Which AETs Have Been Developed in Puget Sound. 

METALS 

Antimony 

Arsenic 

Cadmium 

Chromium 

-pper Nickel 

Lead Silver 

Mercury zinc 

ORGANIC COMPOUNDS 

Low-Molecular-Weight PAHa 

Naphthalene 

Acenaphthyiene 

Acenaphthene 

Ruorene 

Phenanthrene 

Anthracene 

P-Methylnaphthalene 

High-Molecular-Weight PAHs 

Fluoranthene 

Pyrene 

Benz(a)anthracene 

Uvysene 

Benzofluofanthenes 

Benzo(a)pyrene 

Indeno(l,2,3-c,d)pyrene 

Dibenzo(a,h)anthracene 

Benzo(g,h,i)perylene 

Chlorinated Benzeneo 

1.3Dichloroberuene 

1,4-Dichlorobenzene 

1,2DichIoroberuene 

1,2,4-Trichlorobenzene 

Hexachlorobenzene (HCB) 

Phthalates 

Dimethyi phthalate 

Diethyl phthalate 

Di-n-bulyl phthalate 

Duty1 benzyl phthalate 

Bis(2-ethylhexyl)phthalate 

Di-n-octyl phthaiate 

Total PCBe Phermlr 

Phenol 

P-Methylphenol 

4-Methylphenol 

2,4-Dimethylphend 

Pentachlorophenol 

Peatlcldea 

p.p’-DDE 

p,p’-DDD 

p,p’-DDT 

Mirccllrneour Extracta#ea 

Bemyt alcohol 

Benzdc acid 

Dibenzofwan 

Hewchlorobutediane 

N-Nitrosodiphenylamine 

Volatile Orgrnlcs 

Tetrachloroethene 

Ethylbenzene 

Total xylenes 

By normalizing chemical concentrations to appropriate sediment variables (e.g., percent organic 

carbon), differences between different sediment 

types can be minimized in the generation of 

AETs. In practice, identification of unique or 

atypical sediment matrices is important in determining the general applicability of ALIT values 

generated from a specific set of data. 

Differences in physical characteristics (e.g., 

grain size, habitat exposure) are one major factor 

that may amount for stations not meeting predictions based on existing AET values. In Puget 

Sound studies, for example, fine-grainedsediments 

dominated stations that bad significant amphipod 

mortality that had not been predicted, and coarsegrained sediments dominated stations that had 

significant depressions in benthic infauna that had 

not been predicted by benthic AETs (Barrick et 

al., 1988). 

11.3.1.2 Suitability for Different Chnica1.s or 

Classes of Chemicals 

There are no constraints on the types of 

chemicals for which AETs can be developed. An 

AET can be developed for any measured cfremicai 

11-9 
Sediment CZass+i&m Methods Compendium 

(organic or inorganic) that spans a wide-concentration range in the data set used to generate ARTS. 

The availability of a wide diversity of chemical 

data increases the probability that toxic agents (or 

chemicals that covary in the environment with 

toxic agents) can be included in interpreting 

observed biological impacts. 

To date, ARTS have been developed for over 

60 chemicals frquently detected in the environment, including 16 polycyclic aromatic hydrocarbons (PA%); several alkylated PAHs and related 

nitrogen-, sulfur-, and oxygen-containing heterocycles; polychlorinated biphenyls (PCBs) (reported 

as total PCBs); 5 chlorinated benzenes; 6 phthalate 

esters; 3 chlorinated hydrocarbon pesticides; 

phenol and 4 alkyl-substituted and chlorinated 

phenols; 10 metals and metalloids, 3 volatile 

organic compounds; and 5 misce.llaneous extractable substances. Data for other miscellaneous 

chemicals that were less frequently detected or 

analyzed for in the Puget Sound area were also 

evaluated for their potential use in developing 

ARTS (e.g., resin acids and chlorinated phenols in 

selected sediments from areas influenced by pulp 

and paper mill activity). 

ARTS have been developed for chemical 

concentrations normalized to sediment dry weight 

and sediment organic carbon content (expressed as 

percent of dry weight sediment). Using a 188- 

sample data set from Puget Sound, ARTS were 

also developed for data normalized to fine-grained 

particle content (expressed as the percent of silt 

and clay, or <63-urn particulate material, in dry 

weight of sediment). These latter ART values did 

not appear to offer advantages in predictive reliability over the more commonly used dry weight 

and TOC normalizations (Beller ef al., 1986). 

11.3.1.3 Suitability for Predicting Eficts on 

Different Organisms 

The AET approach can be used to predict 

effects on any life stage of any marine or aquatic 

organism for which a biological response to 

chemical toxicity can be determined. Because the 

approach is empirical, relying on direct measurement of the chemical concentrations associated 

with samples exhibiting adverse effects, the results 

are directly applicable to predicting effeds on the 

organisms used to generate the AET. ‘Ihe results 

can also be used to predict effects on nontarget 

organisms by ensuring that the organisms used to 

generate an ART are either representative of the 

nontarget organisms or are more sensitive to 

chemical toxicity than those organisms. For 

example, ARTS generated for a species of sensitive amphipod might be considered as protective 

of the chemical concentrations associated with 

adverse effects in other species of equally or less 

sensitive amphipods. At the same time, these 

ART might be considered protective of most other 

benthic macroinvertebrate taxa because they are 

based on a member of a benthic taxon (i.e., 

Amphipoda) that is considered to be sensitive to 

chemical toxicity (Bellan-Santini, 1980). By 

contrast, AETs generated for a pollution-tolerant 

species such as the polychaete Capitella capitata 

(cf. Pearson and Rosenberg, 1978) might be 

considered representative for other pollutiontolerant species, but not protective for most other 

kinds of benthic macroinvertebrates. 

1 l-3.1.4 Suitability for In-Place Pollutant Control 

In remedial action programs, assessment tools 

such as the ART approach can be used to address 

the following specific regulatory needs: 

Provide a preponderance of evidence for 

narrowing a list of problem chemicals 

measured at a site; 

Provide a predictive tool for cases in 

wbicb site-specific biological testing 

results are not available; 

Enable designation of problem areas 

within the site; 

Rovide a consistent basis on whi& to 

evaluate sediment contamination and to 

separate acceptable from unacceptabk 

COditiOflS; 

Provide an environmental basis for triggering sediment remedial action; and 

II-10 
II-AET Approach 

8 Provide a reference point for establishing 

a cleanup goal. 

Because AET va!ues are derived from sediments 

with multiple contaminants, they incorporate the 

influence of interactive effects in environmental 

samples. The ability to incorporate the influences 

of chemical mixtures, either by design or default, 

is an advantage for the assessment of in-place 

pOllUtantS. 

11.3.1.5 Suitability for Source Control 

The AET approach is well suited for identifying problem areas. Because specific cause-effect 

relationships are not proven for specific chemicals 

and biological effects, remedial actions should not 

be designed exclusively for a specific chemical. 

(This caution applies to all approaches because of 

the complex mixture of contaminants in environmental samples.) The link between problem areas 

and potential sources of contamination is established by analysis of concentration gradients of 

contaminants in these problem areas and the 

presence and composition of contaminants in 

sediments and source materials. The AET approach provides a means of narrowing the list of 

measured chemicals that should be considered for 

source control and provides supportive evidence. 

for eliminating chemicals from consideralion that 

appear to be present at a concentration too low to 

be associated with adverse biological effects. 

Reduction of the overall contaminant load to a 

problem area such that all measured chemicals are 

below their respective AETs is predicted to result 

in mitigation of the adverse biological effects. It 

is possible that such source controls may be 

effective because of the concomitant removal of 

an unmeasured contaminant. 

I I.3. I.6 Suitability for Disposal Applications 

The evaluation of potential biological impacts 

associated with the disposal of dredged material is 

an important component in the designation of 

disposal sites and review of disposal permits for 

dredged material. AET values provide a prepondcrance of evidence in determining a “reason to 

believe” that sediment contamination could result 

in adverse biological effects. Hence, the AET 

approach is a useful tool for assessing the need for 

biological testing during the evaluation of disposal 

alternatives. It is assumed that ABI values 

generated for in-place sediments provide a useful 

prediction of whether adverse biological effects 

will occur in dredged material after disposal at 

aquatic sites. 

113.2 General Advantages and Umitations 

11.3.2.X Ease of Use 

In this section, “use” is treated as both generation and application. The ease of generating AET 

values depends on the status of the data to be used 

for AET generation (i.e., whether field data have 

been collected and whether statistical significance 

has been determined for biological indicators). It 

is recommended that a search for existing data be 

conducted as part of determining the need for 

collecting new samples. The existing database of 

matched biological and chemical data from Puget 

Sound comprises over 300 samples. Collection of 

new field data (e.g., for application outside of 

Puget Sound) would require a considerable expenditure of effort, as would the statistical analysis of 

a large number of samples. However, if data are 

available and statistical analyses have been performed, the generation of APT values is very easy 

with the SEDQUAL database (described in Section 11.2.1.2.4). The menu-driven system allows 

for a considerable amount of flexibility in choosing stations and biological indicators to be included in AET generation. Application of AET (i.e., 

comparison of ABT values to chemical concentrations in field samples) is also very easy when 

using SEDQUAL, provided that the field data 

have been computerized. Application of AEI’ 

values to chemical data presented in existing 

literature is also straightfonvard. 

11.3.2.2 Relative Cost 

The cost of developing ABT values can span 

a wide range, depending on the stage of database 
Sediment CZassif;urtion Methods Compendium 

development and the numbers and kinds of chemicals and biological indicators used. The least 

costly means of developing the values is to use 

existing chemical and biological information, thus 

minimizing the expenses associated with field 

sampling and laboratory analyses. (Selective 

sampling to confirm whether existing AFT vahtes 

are applicabie would still be useful.) The historical database could be based on the pooled results 

from various studies conducted in a region, providing that each study passed QAfQC performance 

criteria and satisfied the prerequisites of the AET 

approach (e.g., matched chemical and biological 

measurements and the ability to discriminate 

adverse biological effects). 

If the historical database is judged inadequate 

to generate AETs for a region, then the costs of 

field measurements of chemical concentrations in 

sediments and associated biological effects must 

be incurred to develop the database. These costs 

can vary substantially, depending on the chemicals 

and biological indicators evaluated. Costs would 

be minimized if evaluations were based on a 

limited range of chemicals and a single, inexpensive biological test. It is recommended that the 

approach be based on a relatively wide range of 

chemicals, and if possible, several kinds of biological indicators. 

The existing database for the Puget Sound 

region is based on a wide range of chemicals (i.e., 

U.S. EPA priority pollutants and other selected 

chemicals) and four kinds of biological indicators. 

The costs for deveIoping AETs varied considerably among the four indicators. For example, 

laboratory costs for the least expensive indicator 

(Microtox bioassay) were approximately $200 per 

station, whereas costs for the most expensive 

indicator (abundances of benthic macroinvertebrates) were as high as $1,800 per station. Tberefore, within the existing database, the range of 

costs for biological testing spanned almost 1 order 

of magnitude. 

Once AET values have been generated, use of 

these values to predict the occurreact of biological 

effects is relatively inexpensive. Chemicat data 

may be compared to AFT values by using the 

SEDQUAL database or through manual data 

manipulations. 

1X.3.2.3 Tendbuy to Be Conservative 

The empirical, field-based nature of &he AFZ 

approach precludes defmitive a priori predictions 

of its tendency to be either over- or underprotective of the environment. The occurrence of 

biologically impacted stations at concentrations 

below the AET of a given chemical (see Figure 

11-1) may appear to be underprotective. However, the occurreo& of impacfed stations at concerttrations below the AFT of a single chemical does 

not imply that AETs in general are not potective 

against biological effects, only that single chemicals may not account for all stations with biological effects. If AETs are developed for multiple 

chemicals, the approach can account for a high 

percentage of stations with adverse biological 

effects. 

To date, AlXs have been developed for acute 

sediment bioassays of mortality in adult amphipocls, developmental abnormality in larval 

bivalves, and metabolic alterations in bacteria. All 

of these organism/endpoint combinations are 

considered to be sensitive to chemical toxicity. 

AETs have also been generated for in sir~ reductions in the abundances of benthic maaoinvertebrates. Because these reductions incaporate chronic (i.e., long-term) exposure to contaminants, they can also be considered as sensitive 

measures of the effects of chemical toxicity. 

However, a more protective approach would be to 

use tbe lowest of the four kinds of AFT for each 

chemical as the concentration on which predictions are made. Alternatively, the protectiveness 

of any kind of AFT could be modified by developing sediment quality values based on “safety 

factors” applied to existing AETs. 

11.3.2.4 Lewd of Acceptance 

The AFT approach has been accepted by 

several federal and state agencies in the Puget 

Sound region as one tool in providing guidelines 

for regulatory decisions. U.S. EPA has used AEI’ 

values to develop sediment quality values with 

which to evaluate the potential toxicity of amtammated sediments in urban bays. PSDDA has used 

AET values as n tc& to develop chemical guide- 
1 I-AET Approach 

lines for determining whether biological testing is 

necessary for dredged sediments proposed for 

unconfined, open-water disposal. Ecology has 

used AET to develop sediment management 

standards. These standards were promulgated by 

the State of Washington and approved by EPA 

Region X in 1991. The standards are being used 

by a number of water quality programs (e.g., 

source control, remediation). 

Several major characteristics influence the 

acceptability of the AET approach. The most 

attractive characteristic of the approach is ptobably the reliance on empirical information based on 

field-collected sediments or indigenous organisms, 

and exposure of laboratory test organisms to 

environmental samples. A second attractive 

feature of the approach is the setting of an AET at 

the chemical concentration in the data set above 

which adverse biological effects are always observed. This characteristic provides consistency 

that, with a representative database used to generate AETs, enhances the preponderance of evidence 

of adverse effects in the environment. The AET 

values can be updated as new information is 

collected The AET approach can also be applied 

to an existing database in new regions, providing 

certain prerequisites are met by the database (e.g., 

synoptic measurement of chemical and biological 

data, and QA/QC guidelines). 

A limitation of the AET approach is that 

field-based approaches do not directly assess 

cause-effect relationships. Because sediments in 

the environment are often contaminated with a 

complex mixture of chemicals, it is difficult when 

using field-collected sediment for any approach to 

relate observed biological effects to a single 

chemical. The approach also requires selection of 

appropriate normalized chemical data to address 

the bioavailability of contaminants to organisms. 

Organic carbon notmalization may be most appropriate for nonpolar organic contaminants based on 

theoretical considerations. In addition, nonprotective AETs could be generated if unusual matrices 

(e.g., slag) that anomalously restrict bioavailability 

are included in the database used to generate the 

AETs, or if biological test results are incorrectly 

classified. Recommended data treatment guidelines for chemical and biological data are discussed by Barrick et al. (1988). The AET ap 

preach was reviewed by the U.S. EPA Science 

Advisory Board (SAB, 1989), which noted the 

method had “major strengths in its ability to 

determine biological effects and assess interactive 

chemical effects.” 

11.3.2.5 Ability to Be Implemented by 

Laboratories with 7’ypical EquipmeW 

and Handling Facilities 

If applicable data do not already exist, the 

development of AET values requires a relatively 

extensive amount of field sampling and laboratory 

analysis. The chemical analyses required for 

development of AET represent standard analytical 

procedures. A laboratory with appropriately 

trained staff should be able to conduct the necessary benthic community analyses and sediment 

bioassays. Specific methods for performing the 

chemical and biological tests that were used to 

develop Puget Sound ABT are detailed in the 

Puget Sound Protocols (Tetra Tech, 1986). These 

efforts can be minimized by using historical data 

whenever posstble. Once AETs are developed, 

their routine implementation is relatively easy. In 

addition, they can be easily updated as additional 

data become available. 

11.3.2.6 Level of Effort Required to Generate 

ReWh 

As noted in Section 11.3.2.1, the SEDQUAL 

database facilitates AET generation and application. After field data have been collected, the 

most time-consuming task is data entry and 

verification. Entry of chemical and biological data 

for 50 samples requires roughly 16 person-hours 

(assuming 75 chemicals have been measured and 

biological effects are being coded simply as 

“impacted” ot “nonimpaded”). Generating a set 

of AET values for a given biological indicator, 75 

chemicals, and 50 stations takes approximately 

0.751 h of computer time on SEDQUAL (and 

about 5 min of labor to set up the analysis). To 

compare a set of AET (for 75 chemicals) to a SOsample set of field data takes approximately 0.5 

0.75 h of computer time on SEDQUAL (and 

21-13 
Sedimmt Ciassi&&on Methods Compendium 

roughly 5 mia of labor to set up the analysis). 

SEDQUAL is capable of comparing any kind of 

chemical sediment aiteria to field data, but requires that the numerical criteria be entered in the 

database. 

11.3.2.7 Degree to which Results Lend 

Themselves to Interpretation 

The manner in which the AET approach can 

be used to interpret matched biological and chemical data from field-collected sediments is described in Section 112.1. As noted previously, 

the use of AET can help investigators eliminate 

chemicals from further consideration (as the cause 

of an observed effect); however, the approach 

cannot identify specific cause-effect relationships. 

Because the AET approach is empirical, it is not 

well suited to identifying specitic toxic agents or 

elucidating mechanisms of biological uptake and 

metabolism. However, certain general relalionships could be examined on an a posteriori basis 

with the AET approach (e.g., testing the relative 

importance of different ways of normalizing 

chemical concentration data in predicting adverse 

biological effects). 

A number of environmental factors may 

complicate the interpretation of the data. Although the AET concept is simple, the generation 

of AET values based on environmental data 

incorporates many complex biological-chemical 

interrelationships. For example, the AET approach incorporates the net effects of the following factors that may be important in fieldcollected 

sediments: 

n Interactive effects of chemicals (e.g., 

synergism, antagonism, and additivity); 

n Unmeasured chemicals and other unmeasured, potentially adverse variables; and 

w Matrix effects and bioavailability (i.e., 

phase associations between contaminants 

and sediments that affect bioavailability of 

the contaminants, such as the incorporation of PAH in soot particles). 

‘fbe AET approach cannot quantify the individual contributions of interactive effects, unmeasured chemicals, or matrix effects in environmental samples, but AET values may be influenced by 

these factors. AET values are expected to be 

reliable predictors of adverse effects that could 

result from the influence of these environmental 

factors if the samples used to generate AETs are 

representative of samples for which AET predictions are made. Alternatively, isolated occurrences of such environmental factors in a data set used 

to generate AETs may limit the predictive reliability of those AET values. If confounding environmental factors render the AET approach unreliable, then this should be evident from validation 

tests in which biological effects are predicted in 

adual environmental samples. 

A more detailed discussion of the interpretation of AETs and the confounding effects of 

environmental factors is presented in U.S. EPA 

(1988). 

11.3.2.8 Degree of Environmental Applicability 

The AEIT approach has a high degree of 

environmental applicability based on its reliance 

on chemical and biological measurements made 

directly on environmental samples. Such information provides tangible evidence that various 

chemical conceatralions either are or are not 

associated with adverse biological effects in 

typically complex environmental settings. 

The environmental applicability of the AET 

approach has been quantified for the four kinds of 

AET developed for Puget Sound by evaluating the 

reliability with which each kind of AET predicted 

the presence or absence of adverse biological 

effects in field samples collected from Puget 

Sound (USEPA, 1988). The overall reliabiiity of 

the four tests ranged from 85 to 96 percent, 

indicating that all four kinds of AITs were relatively accurate at predicting the prez3ence or 

absence of effects for samples from the existing 

database. This high level of reliability suggests 

that AETs have a relatively high degree of environmental applicability in Puget Sound, and it has 

been a primary factor in the use of the AET’ 

approach by agencies in the Puget Sound ~&OIL 

AET values generated for Puget Sound have also 

been used as examples of effects-based sediment 

11-74 
II-AET Appach 

criteria to provide an initial estimate of the magnitude of potential problem areas in coastal regions 

of the United States for the U.S. EPA Office of 

Policy Analysis (PTI, 1987). 

11.3.2.9 Degree of Accuracy and Precision 

In this section, accuracy is considered to be 

the ability of AET to predict biological effects and 

precision represents the expected variability 

(uncertainty range) for a given AET value for a 

given data set. 

In previous evaluations of the AET approach 

and other sediment quality values using fieldcollected data, the accuracy of the approach was 

defined by two qualities: 

l Sensitivity in detecting environmental 

problems (i.e., are all biologically impacted sediments identified by the predictions 

of the chemical sediment criteria?) 

l Efficiency in screening environmental 

problems (i.e., are only biologically impacted sediments identified by the predictions of the chemical sediment criteria?). 

Sensitivity is defined as the proportion of all 

stations exhibiting adverse biological effects that 

are correctly predicted using sediment criteria. 

Efficiency is defined as the proportion of all 

stations predicted to have adverse biological 

effects that actually are impacted. Ideally, a 

sediment criteria approach should be efftcient as 

well as sensitive. For example, a sediment criteria 

approach that sets values for a wide range of 

chemicals near their analytical detection limits will 

likely be conservative (i.e., sensitive) but inefficient. That is, it will predict a large percentage of 

sediments with biological effects. It will also 

predict impacts at many stations where there are 

no biological effects, but chemical concentrations 

are slightly elevated. The concepts of sensitivity 

and efficiency are illustrated in Figure 11-2. 

The overall reliability of any sediment criteria 

approach addresses both sensitivity and efficiency. 

This measure is defined as the proportion of all 

stations for which correct predictions were made 

for either the presence or absence of adverse 

biological effects: 

High reliability results from correct prediction of 

a large percentage of the impacted stations (i.e., 

high sensitivity, few false negatives) and correct 

prediction of a large percentage of the nonimpacted stalions (i.e., high efficiency, few false 

positives). An assessment of AET reliability was 

recently conducted using a large database comprising samples from 13 Fuget Sound embayments 

(Barrick et al., 1988). These evaluations suggest 

that the AET approach is relatively sensitive for 

the biological indicators tested and also relatively 

efficient. For example, 68-83 percent sensitivity 

and 55-75 percent efficiency were observed when 

AETs generated from a 188-sample data set were 

evaluated with an independent 146-sample data 

set. The ranges of sensitivity and efficiency cited 

above represent the ability of benthic infaunal 

AET values to predict statistically significant 

depressions in the abundances of benthic infauna 

in field-collected samples and the ability of amphipod mortality bioassay AET values to predict 

statistically significant mortality in bioassays 

conducted on field-collected sediment. 

Precision of the AET approach has not been 

as intensively investigated as accuracy. AET 

values are the result of parametric statistical 

procedures (i.e., determination of the significance 

of biological effects relative to reference conditions) and nonparametric methods (e.g., ranking of 

stations by concentration), and thus are not amenable to the routine definition of confidence intervals. However, the degree of AET precision is 

considered to depend on the following factors: 

n ‘Ibe concentration range between the AEl” 

(determined by a nonimpacted station) 

and the next highest concentration that is 

associated with a statistically significant 

effed; 

I aassification error associated with the 

statistical significance of biological indi21-15 
S&nent Chassftition Methods Compendium 

IMPACTED 

CORREClLY PREDICTED 

PREDICTED 

SENSITIVITY = C/B x 100 = 5/8 x 100 = 63% 

EFFICIENCY = C/A x 100 = 5/7 x 100 = 71% 

FOR A GIVEN BIOLOGICAL INDIMTOR: 

A AK STATIONS PREDICTED TO BE IMPACTED 

B ALL STATIONS KNOWN TO BE IMPACTED 

C AU STATIONS CORRECTLY PREDICTED TO BE IMPACTED 

Figure 1 l-2. Measures of reliability (sensitivity and efhlency). 

cator results (i.e., whether a station is 

properly classified as impacted or nonimpacted, as related to Type I and Type II 

statistical error); 

m The weight of evidence or number of 

observations supporting a given AET 

value; and 

l Tke analytical error associated with quantification of chemical results. 

Detailed discussion of these factors is provided in 

Beller et al. (1986). 

One approach used in Puget Sound to estimate 

the uncertainty range around the AET value was 

to define the lower limit as the concentration at 

the nonimpacted station immediately below the 

AT3 and to define the upper limit as the concentration at the impacted station immediately above 

the AET. These limits are based largely on 

probabilities of statistical classification error. For 

data sets with large concentration gaps between 

stations, such uncertainty ranges will be wider and 

precision will be poorer than for data sets titb 

more continuous distributions. ‘Ike number of 
1 I-AET Approach 

stations used to establish an AET would be expected to have a marked effect on AET uncertainty because small data sets would tend to have less 

coalinuous distributions of chemical concentrations than large data sets. Based on analyses 

conducted with Puget Sound data, the magnitude 

of the AET uncertainty for 10 chemicals or chemical groups that are commonty detected is typically 

less than one-third to one-half of the value of the 

AET itself (considering botb amphipod mortality 

bioassay and benthic infaunal AET data). Based 

on quality assurance information for these data, 

analytical error is probably a minor component of 

overall precision, particularly for metals. 

11.4 flA?X'Us 

11.4.1 Extent of Use 

The AET approach is used by several agencies 

and sediment management programs in the Pacific 

Northwest to provide guideline values for regulatory decisions. The State of Washington has 

developed sediment management standards primarily using the AET approach but also including 

equilibrium ptiitioning values. These standards 

were promulgated by the State and approved by 

EPA, Region X, in 1991 and are currently being 

implemented in a variety of programs. The 

standards are the culmination of cooperative 

planning and scientific investigations by several 

federal and stale agencies throughout the 1980’s, 

including: 

m Superfund investigations at Commencement Bay and Eagle Harbor; 

m Puget Sound Dredged Disposal Analysis 

(PSDDA); 

l Urban Bay Toxics Action Program; and 

m Puget Sound Water Quality Authority 

Management Plan. 

A key result of these efforts has been the rewgnition by regulators of two separate levels of sediment contamination and has led to tie deuelopment of two sets of sediment quality values. This 

separation in management use of sediment values 

arose from the sensitivity and efficiency concepts 

of reliability previously discussed. This management decision was made because it was determined that none of the available approaches for 

developing sediment quality values would result in 

100 percent sensitive and 100 percent efficient 

values. Different strategies have been used by 

different programs for use of AET-generated 

values. In general, the lowest AET (termed 

LAET) for any of the biological tests is used to 

establish the lower level where there is little 

concern of se&men1 contamination (e.g., Ibe goal 

for remedial actions). The AET approach has 

developed higher chemical levels (termed m, 

above which adverse effects are predicted for all 

the biological tests. In most regulatory programs, 

direct biological testing is allowed to resolve the 

differences in predictions of t&e two sets of 

sediment quality values (Le., prediction of adverse 

biological effect by highly sensitive sediment 

quality values, which at lower chemical concentrations are not predicted by highly efficient sediment quality values). To date, such sediment 

quality values developed were for and used in 

marine and estuarine environments. The State of 

Washington and EPA, Region X, are gathering 

chemical and biological data to potentially develop 

companion values for freshwater sediments. 

Other efforts are under way outside Puget 

Sound and the Pacific Northwest to develop sediment quality values using the AET approach. mese 

include California and the Great Lakes region ia tbe 

United States, and the countries of Canada, New 

Zealand, and Australia internationally. 

11.4.2 Extent to W&b Approach Has Been 

Field-VJidnted 

As described in U.S. EPA (1988), the reliability of AETs generated from Puget Sound data WBS 

evaluated with tests of sensjtivily aad efficieacy 

(defined ia Section 11.329). Tests of tie seusitivity and efficiency of the AET appsoa& were 

carried out in several steps, as deskbed below: 
Sediment Classficution Methods Contpmdium 

l The chemic8I database was subdivided 

into groups of stations that were tested for 

the Same biologic81 effects indicators. 

Specifically, ali chemistry stations with 

associated amphipod bioassay data were 

grouped together (287 stations), all chemistry stations with associated benthic 

infaunal data were grouped together (201 

stations), all chemistry stations with associated oyster larvae bioassay data were 

grouped together (56 stations), and all 

chemistry stations with associated Micro 

tox bioassay data were grouped together 

(50 stations). Stations with more than 

one biologic81 indicator were included in 

each appropriate group. 

n The stations in each group were classified 

8s impacted or nonimpacted based on the 

appropriate statistical criteria (i.e., F, 

and t-tests at alpha = 0.05). 

n Several tests of reliability were conducted 

at this point: 

l Test 1: AET values (dry weight) 

were generated with the entire 

l+tget Sound database available in 

1988, and sensitivity and efficiency 

tests were performed against the 

same database for each biological 

indicator. 

l Test 2: The test described above 

WaS repeated in hV0 park: (8) Using 

T0C-IlO~81iZd liJ?f V81UfZ.S for 

nonionic organic compounds and 

dry weight-normalized AET values 

for all other compounds (i.e., ionizable organic compounds, metals, 

and metalloids), and (b) using 

mC-norm8bzed data for 811 chemicals. Test 2 allowed for oposteriori evaluation of the relative success 

of dry weight and TOC normaiization for nonionic organic chemicals. 

l Test 3: Because the efficiency of 

tie AET based on the entire Puget 

Sound database is 100 percent by 

constraint (as in Tests 1 and 2), 

predictive efficiency was e8timated 

by the following procedure. For 

each biological indicator, a single 

station WBS sequentially deleted 

from the total database, AETs were 

recalculated for the reumining data 

set, and biologic81 effects were 

predicted for the single deleted 

station. ‘Zbe predictive efficiency 

was the cumulative result for the 

sequential deletions of single stations. For example, the 287~sample 

database for amphipod bioassay 

results cztn be used to provide a 

286~sample independent database 

for predicting (in sequence) effects 

on 811 287 samples. 

l Test 4: In this test, independent 

data sets were used to generate and 

test AETs to confirm the sensitivity 

and efficiency me8suremenk in 

Tests 1 and 3. AETs (dry weight) 

generated with 188 stations from 

diverse geographic regions in Puget 

Sound were tested with a comple 

tely independent set of 146 Fuget 

Sound stations. 

In addition, the influence of geographic 

location and other factors on AET predictive 

ability were examined (Banick et ol., 1988). 

Further testing of Pttget Sound AET v8hes using 

matched biological/chemical data from other 

geographic areas is desirable before recommending dire& application of the FItget Sound values in 

other geographic regions. 

11.43 Reasons for Limited Utw 

The AET approach is being inue8singly used 

outside of Puget Sound and the Pacific Northwest 

to evaluate and compare different classes of 

sediments and to develop bay-, site-, or regionspecific sediment quality values for a variety of 

regulatory uses. Because the approach is based on 

empirical data, direct application of values from 
II-AET Approach 

Puget Sound or another area to a specific bay, 

site, or region usually encounters some conflicting 

or confounding data. Because regional reference 

areaS are used to determine the significance of 

adverse biological effeck in the AET approach, 

the AET developed for one region may be overprotective or underprotective of the resources in 

the other area. Additionally, the mix of chemicals 

in one region’s sediments may not be the same in 

another region. The use of the AFZT approach 

and use of specific AET values should not be confused. 

Development of site-specific AETs for other 

geographic areas may require additional sampling. 

Because many past studies were not multidisciplinary, measurements were often made only for 

chemistry or biology rather than for both kinds of 

information. In such cases, there will be 8 limited 

amount of appropriate historical data that can be 

used to develop AETs. The integration or comparison of AET data sets among different regions 

can also be restricted because appropriate biological indicators for generating AETs may vary 

among regions. 

11.4.4 Outlook for Future Use 8nd Amount of 

Development Yet Needed 

The following two approaches to AET development could be particularly beneficial in expanding the use of this approach: 

n Use of laboratory cause-effect (spiking) 

studies to evaluate AET predictions on a 

chemical-specific basis and 

n Use of a large set of matched biological/ 

chemical data from different geographic 

areas to teat the predictive ability of AET 

and to test the “precision” of AET values 

based on data sets from different areas. 

The AET approach was presented to (USEPA, 

1988) and reviewed by tbe U.S. EPA Science 

Advisory Board (SAB, 1989). The SAB noted 

major strengths and limitations of tie method and 

provided recommendations that would improve the 

validity of the AET values. The method was 

considered to contain sufficient merit for use in 

developing location-specific sediment quality 

values. Because of the specificity of the method, 

i.e., the empirical applications at specific localities, under specific environmental conditions, the 

approach seemed less useful for development of 

general, broadly applicable (i.e., national) sediment quality criteria. 

11.5 REFERENCES 

Barrick, R.C., S. Becker, L Brown, H. Beller, and 

R. Pastorok. 1988. Sediment quality value8 

refinement: 1988 update and evaluation of 

Puget Sound AET. Volume I. Final Report. 

Prepared for Tetra Tech, Inc. and U.S. Environmental Protection Agency Region X, Office of 

Puget Sound. PI’I Environmental Services, 

Bellevue, WA. 74 pp. + appendices. 

Becker, D.S., R.P. Pastorok, R.C. Barrick, P-N. 

Booth, and LA. Jacobs. 1989. Contaminated 

sediments criteria report. Prep8red for the 

Washington Department of Ecology, Sediment 

Management Unit. PI’I Environment81 Services, Bellevue, WA 99 pp. + appendices. 

Bellan-Santini, D. 1980. Relationship between 

populations of amphipods and pollution. Mar. 

Poll. Bull. 11224227. 

Beller, H-R., R.C. Barrick, and D.S. Becker. 

1986. Development of sediment quality values 

for Puget Sound. Prepared for Resource Planning Associates, U.S. Army Corps of Engineets, 

Seattle District, and Puget Sound Dredged 

Disposal Analysis Program. Tetm Tech, Inr, 

Bellevue WA. 128 pp. + appendices. 

Nielsen, D. 1988. SEDQUAL users manual. 

Prepared for Tetra Tech, Inc. and U.S. Environmental Protection Agency Region X, Office of 

Puget Sound. PII Environmental Services, 

Bellevue, WA. 

Pearson, T.H., and R. Rosenberg. 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine enviroament. Oceanogr. Mar. Biol. Annu. Rev. 16: 

229-311. 

11-19 
Sediment Classification Methods Compendium 

Phillips, K., P. Jamison, 1. Malek, B. Ross, C. 

Krueger, J. Thornton, and J. Krull. 1988. 

Evalualion procedures technical appendixPhase 1 (Central Puget Sound). Prepared for 

Puget Sound Dredged Disposal Analysis by the 

Evaluation Procedures Work Group. U.S. 

Army Corps of Engineers, Seattle, WA. 

Puget Sound Water Quality Authority. 1988. 

1989 Puget Sound Water Quality Management 

Plan. Puget Sound Water Quality Authority, 

WA. 276 pp. 

PTI. 1987. Policy implications of effects-based 

marine sediment criteria. Prepared for American Management Systems and U.S. Environmental Protection Agency, Of!% of Policy 

Analysis. PTI Environmental Services, Bellevue, WA. 

PTI. 1988. Elliott Bay Action Program: 1988 

action plan. Prepared for Telra Tech, Inc. and 

U.S. Environmental Protection Agency. PTI 

Environmental Services, Bellevue, WA. 43 pp, 

+ appendices. 

Sokal, R.R., and FJ. Rohlf. 1%9. Biometry. 

W.H. Freeman and Company, San Francisco, 

CA. 859 pp. 

State of Wasbington, Department of Ecology. 

1991. Chapter 173-204, Washington Administrative Code, Sediment Management Standards. 

Olympia, WA. 

Swartz, R.C., WA. DeBen, J.K Phillips, J-D. 

Lunberson, and FA. Cole. 1985. Phoxocephalid amphipod bioassay for marine sedimenl 

toxicity. pp. 284-307. In: Aquatic Toxicology 

and Hazard Assessment: Proceedings of the 

Seventh Annual Symposium. R.D. Cardwell, R. 

Purdy, and R.C. Bahner (eds.). ASTM STP 854. 

American Society for Testing and Materials, 

Philadelphia, PA. 

Tetra Tech. 1986. Recommended protocols for 

measuring select%4 environmenta variables in 

Puget Sound. Fina! report. Prepared for U.S. 

Environmental Protection Agency, Region X, 

Office of Puget Sound, Seattle, WA. Tetra 

Tech, Inc., Bellevue, WA. 

USEPA, 1989. Science Advisory Board. Report 

of the Sediment Criteria Subcommittee, Evaluation of the Apparent Effects T&&old (AET) 

Approach for Assessing Sediment Quality. 

SAB-EETFC-89-027. Office of the Administrator, Science Advisory Board, Washington, DC. 

USEPA. 1988. Briefing report to the EPA Science Advisory Board. Prepared for Battelle and 

U.S. Environmental Protection Agency, Region X, Office of Puget Sound. PT’I Environmental Services, Bellevue, WA. 57 pp. 
CHAPTER 12 

A Summary of the Sediment Assessment 

Strategy Recommended by the International 

Joint Commission 

Philippe Ross 

The Citadel, Department of Biology 

Charleston, SC 29409 

(803) 792-7875 

The International Joint Commission (IJC) 

Sediment Subcommittee has published a document 

entitled Procedures for the Assessment of Contaminated Sediment Problems in the Great Lakes (IJC, 

1988a). An overview of the IJC strategy for 

assessing contaminated sediments is provided in 

this chapter. However, because it would be 

inappropriate to reproduce all, or substantially all, 

of the document in this chapter, the interested 

reader is referred to the IJC (1988a) document 

itself for an explanation of details that are not 

provided herein. 

12.1 SPECIFIC APPLICATIONS 

12.1.1 Current Use 

The IJC (1988a) document is intended as 

guidance for the assessment of contaminated 

sediments in the Great Lakes. Its first application 

is in a work plan for sediment investigations at 

Great Lakes areas of concern (AOCs, as identified 

by the IJC). Section 118(c)(3) of the Water 

Quality Act of 1987 calls for U.S. EPA’s Great 

Lakes National Program Office to survey at least 

five AOCs as part of a 5-yr study and demonstration program called ARCS (Assessment and 

Remediation of Contaminated Sediments). The 

strategy recommended by IJC (1988a) will be 

applied through a series of activities involving 

physical mapping and characterization, sampling, 

chemical analyses, toxicity testing, and in situ 

community analysis. The assessment began in 

1989 and was completed in 1991. The ARCS 

program also seeks to improve upon the IJC 

(1988a) approach by comparing various test 

methods and by evaluating cost-effective reconnaissance and screening methods. 

12.1.2 Potential Use 

Other AOCs will eventually be evaluated in 

the process of developing remedial action plans. 

It is possible that other Great Lakes harbors, 

rivers, and estuaries will be added to the list of 

AOCs, in which case remedial action plans would 

have to be developed there. In addition, the 

guidance document could potentially be used to 

assess suspected sediment contamination outside 

the Great Lakes basin. 

12.2 DESCRIPTION 

12.2.1 Description of Method 

12.2.1.1 Objectives and Assumptions 

In response to the need for a common 

approach to the assessment of contaminated 

sediments, the IJC’s Sediment Subcommittee has 

developed a strategy based on protocols that 

emphasize biological monitoring. The approach 

is intended for use in comprehensive assessments of areas (e.g., bays, harbors, rivers, other 

depositional zones) where sediment contamination and the need for remedial action are suspected. While the suggested strategy attempts to 

minimize the cost and expertise, the assessments 

are relatively large undertakings appropriate to 

situations where large-scale remedial actions 

might be contemplated. In such cases, the cost 
Sediment Classification Methods Compendium 

of conducting accurate assessments would be 

justified if the subsequent remedial options 

could cost far more than the assessments. It was 

not the primary intent of the subcommittee to 

provide guidance for small-scale decision-making activities, such as sample-by-sample disposal 

of dredged material from navigation channels. 

Nevertheless, some of the component methods 

described could be useful and cost-effective in 

this regard. The first major assumption, therefore, is that the scope of the study in question is 

sufficient to warrant a large-scale integrated 

investigation. 

Another fundamental assumption is that the 

ultimate concern of a problem assessment focuses on whether sediment contaminants are exerting biological stress or are being bioaccumulated. Accepting this assumption, it follows that 

adequate assessments of sediment quality should 

involve components of chemistry, toxicity, and 

infaunal community structure (Chapman and 

Long 1983), a concept frequently referred to as 

the Sediment Quality Triad approach (see Chapter 9). The proposed strategy has the following 

objectives: 

To provide accurate assessments of specific problems by using a modified 

‘triad” approach, which integrates chemical, physical, and biological information; 

To perform tasks in a sequence 

the results from each technique 

used to reduce subsequent 

ling requirements and costs; 

To provide adequate proof of 

between the contamination and 

served biological impact; 

To quantify problem severity, 

so that 

can be 

samplinkage 

the obthereby 

enabling intercomparisons between and 

within areas of investigation (thus allowing development of a priority list for 

remedial actions and the objective selection of appropriate remedial options); 

• To consider the effects on different species 

and different trophic levels, since biological 

impairment may occur in the water column 

and the sediments if resuspension occurs 

and since there is no such thing as the 

universal “most-sensitive species” (Cairns, 

1986). 

The IJC approach is an integrated strategy that 

provides the necessary data to identify sedimentassociated contamination as the problem source, 

specify effects, rank problem severity, and assist in 

the selection of remedial options. While the assessment portion of the document identifies a set of the 

best currently available assessment tools (see Section 

l2.2.1.2.2), it is assumed that decisions will be made 

based at the circumstances unique to each AOC 

There is no substitute for experience (expert judgment), and it is also assumed that appropriate 

expertise will be assembled before the assessment 

study plan is formulated. 

12.2.1.2 Level of Effort 

12.2.1.2.1 Type of Sampling Required 

The IJC (1988a) approach involves two stages. 

Stage I, the initial assessment, isused for areas 

where an inadequate or outdated databasexists, In 

the IJC document Stage I is not subdivided, while 

Stage II is broken into Phases I, II, III, and IV. 

Stage I uses only in situ assessment techniques and 

criteria: a limited physical description of the area 

(e.g., basin size and shape, bathymetry) and the 

sediments, bulk chemical analyses, resident benthic 

community organization (e.g., family-level identifications), fish contaminant body burdens (one important species, selected by expert judgment), and 

external abnormalities on collected specimens. Any 

one of the following criteria provides sufficient 

justification for proceeding to Stage II: 

• Concentrations ofmetals above background 

levels in sediments; 

• Concentrations of hazardous persistent 

organic compounds above best available 

detection levels in sediments; 

12-2 
Concentrations of hazardous persistent 

organic compounds above detection levels 

in ftsh or benthos; 

The absence of a healthy benthic community (e.g., absence of clean water organisms 

such as amphipods or mayflies, presence of 

a community dominated by oligochaetes, 

the complete absence of invertebrates); and 

Presence of extental abnormalities in f&h. 

These conditions must be supported by evidence 

that the observed situation is not due to a major 

sediment perturbation, such as dredging or substrate 

modilication. 

Available data may preclude the need for a 

Stage I assessment. The cost and effort that Stage I 

entails should be avoided if there is already strong 

evidence of a contamination problem. 

When a probable sediment contamination 

problem is identified, either through the initial 

assessment or from the examination of existing data, 

then Stage 11, the detailed assessment, should be 

undertaken. llbe detailed assessment consists of 

four phases, which together define the sediment 

problem in the most costeffective manner. The 

phases are not inflexible protocols, but rather logical 

groupings of work units. The expert invesligator 

should be responsible for the final study design. 

ln Phase I of Stage Il, extensive information on 

the physical composition of the sediments is collected. These data are used to define areas or zones of 

homogeneity within a study area. Knowledge of 

these zones allows sampling requirements for 

Phase II to be estimated. 

In Phase II of Stage II, the benthic community 

structure is examined to the lowest possible taxonomic level (e.g., species or variety), along with the 

surlicial sediment chemistry (e.g., pH, total organic 

carbon, redox potential, metals, extractable organic 

compounds). Phase Il results can be combined with 

Phase 1 data to reduce the sampling effort in the 

next phase. 

In Phase Ill of Stage II, a battery of laboratory 

bioassays (e.g., Miuotox, algal, daphnid, benthic 

invertebrate, fii, Ames test) are performed on a 

smaller number of sediment samples than those in 

the Phase Ii sample set. Since fresh sediment must 

be collected for this phase, precision position-f-g 

equipment is required to relocate previously sampled 

sites. Phase III costs can be reduced by performing 

acute lethality bioassays on a sediment sample 

before proceeding to tests that measure chronic or 

sublethal effects. Also in Phase III, sediment cores 

are wlleded, dated, and sectioned for stratified 

chemical analyses and bioassays. Finally, adult fish 

are examined histopathologically for internal (e.g., 

liver) tumors. In relatively confined geographical 

areas, Phases Ii and III may be combined becam 

further sampling may be more ccstly than conduding additional bioassays and relocating Phase Il sets 

for Phase III sampling may be difficult. In this 

case, phase II sampling will include extra material 

for Phase m. 

In the fourth and fmal phase of Stage II sediment dynamics (e.g., accumulation, resuspension, 

movement) and factors affecting them are quantilied. All of the foregoing information is necessary 

for the selection of appropriate remedial options 

For example, depositional history, as revealed by 

sampling sediment cores, and sediment dynamicsare 

critical pieces of information in the selection and 

cost evaluation of remedial options. 

Oitexia that clearly indicate when some form of 

remedial action must be considered (based on the 

results of Stage II) am essential. Because of the 

absence of definitive sediment action criteria at time 

of writing, the criteria proposed by the UC (19&3a) 

are highly conservative, following the language of 

the 1978 Great Lakes Water Quality Agreement as 

tevised in 1987 (especially Annexes 1 and 12), in 

order to promote maximum protection and effective 

restoration of the Great Lakes ecosystem. The UC 

(1988a) urges that these criteria be reviewed regulady to ensure that they continue to fulfill their 

intendedputpose. 

12212.2 Methods 

During Stage & the minimum amount of information necessary to amess potential problem sediments is collected. A variety of physical, chemical, 

and biological measurements are recommended, as 

outlined below 

12-3 
Sediment Classification Methods Compendium 

n 

A geographical description of the area and 

its bathymetry is required. 

Sediment grain size - Size analysis tecbniques based on settling velocity (American Society for Testing and Materials, 

1964; Duncan and IaHaie, 1979) are recommended. The sand fraction is removed by a 62-,um sieve and analyzed 

separately from the fine-grained material. 

Sediment water content - The water content 

can be determined during sample preparation for grain size and other analyses by 

comparison of sample weights before and 

after either freeze-drying or oven-drying 

(Adams et al., 1980). 

Redox potential (Eh) and pH should be 

measured [specific methods are not recommended by IJC (1988a)J. 

Organic carbon - It is recommended that 

total sediment organic carbon be measured 

as descrii by Plumb (1981). 

phosphorus - tie measurements are suggested: total phosphorus, as extracted from 

sediment by sodium carbonate fusion or by 

perchloric acid digestion, and bioavailable 

phosphorus, as estimated by NaOH extractable phosphorus (Williams et al., 1980). 

Ten metals (lead, nickel, copper, zinc, 

cadmium, chromium, iron, manganese, 

mercury, and arsenic) are recommended for 

routine analysis at Great Iakes AOCs. 

Additional metal analyses are left to the 

judgment of the investigator. &I extraction 

procedure using a mix of hydrochloric and 

nitric acids (1: 1) is suggested (Plumb, 

198 1). 

Persistent organic compounds - The reader 

is referred to the U.S. EPA (1984) protocols for broad scans and analyses of individual compounds. When the strategy was 

written, no standardized chemical protocols 

for estimating bioavailability of trace organic compounds were identified. 

l External abnormalities in ftih - ‘Ihe presence of one or more external abnormalities 

is often indicative of anthropogenically 

induced stress or damage. In the case of 

the brown bullhead, Ictakcrus nebukwus, 

phenomena such as stubbed barbels, skin 

discoloration (melanoma), and skin tumors 

are highly correlated with liver cancer 

incidence (Smith et oi., 1988). It is recommended that locally occuning Gall (particularly I. nebukxw) be examined for 

tumors, melanoma, blindneq and barbel 

abnormalities during a Stage I assessment. 

n Contaminant body burdens - The bent& 

infauna are in continuous contact with the 

sediments, providing a direct measure of 

the specific relationship between local&d 

sediment contaminant concentrations and 

bioavaiIability. Cup are also regularly in 

contact with and ingest large quantities of 

sediments. Tbey represent a larger spatial 

and temporal integration of contaminants 

than do the benthic infauna. Colledion of 

adult common carp (Cyprinur carpi for 

tissue residue analysis is recommended. 

Three to five fEb per replicate should be 

cornposited. The number of replicates is 

determined using variability estimates from 

monitoring programs (Schmitt et al., 1983) 

and a chosen level of precision, to calculate 

an idealized sample size @. 247, Sokal and 

Rohlf, 1%9). It is also recommended that 

the most abundant benthic invertebrate 

species (often oligochaete worms in amtaminated sediments) be sampled in early 

summa, prior to thermal stratification. 

Standard U.S. EPA methods are suggested 

for tissue residue analysis. The problem of 

obtaining enough biomass for analysis (at 

least1g)istecognized. 

l Benthic community structure - In a 

Stage I assessment, a preliminary analysis 

of community structure impairment is 

12-4 
recommended. A qualitative study with 

minimal replication and identification only 

to the family level is suggested. Because 

it is important that rare taxa be sampled, 

simple techniques tbat employ inexpensive equipment but take large samples are 

recommended. This approach should 

suffice to identify the existence of a 

stressed community for the purposes of 

Stage I criteria (see Section 12.2.1.2.1 

above). 

Phase II of the detailed assessment consists of 

more focused analyses to supplement or complement 

information obtained in Stage I. Phase I of the 

detailed assessment focuses on physical mapping of 

the environment. Ihe most important aspect of the 

physical assessment of a suspected contaminated 

sediment deposit is its threediiensional mapping. 

A rectangular grid pattern is rt-zommendecl for the 

initial mapping operation. Concurrent with bottom 

sampling at grid intersections, echo-sounder and 

side-scan sonar surveys should be performed to 

improve spatial resolution of sediment zones and 

bottom features. Detailed surveys should include 

piston coring for stratigraphic resolution. The grid 

sampling results should be examined using cluster 

analysis (or similar techniques), which are easy to 

interpret and functional with a small number of 

variables. Basic information required in this phase 

includes geographic location, area1 extent, thickness 

and total sediment volume, average depths of 

overlying water, and the grain size properties of the 

deposit. Phase 1 results are used to select sampling 

sites for later phases. 

Phase II of the detailed assessment focuses on 

surficial sediment chemistry and benthic community 

structure. Based on the previous mapping of homogeneous zones (Phase I), effort in Wase II can be 

expended in depositional areas and in those areas 

with fine-grained sediments. Surflcial chemistry 

sampling should be coincident with the sampling for 

detailed benthic community structure analysis. Total 

organic carbon, redox potential, pH, metals, and 

persistent organics should be measured. Investigators are referred to Plumb (Ml), Williams et OL 

(1980), and U.S. EPA (1984) for collection and 

analysis methods. 

Since the main objective of Stage Il community structure assessment is to examine subtle 

distinctions in stress response, more detailed 

taxonomic data are required in this phase than 

were required in Stage I. In the study design and 

sample collection steps, investigators are urged to 

follow lhe 10 principles of sampling set forth by 

Green (1979). Further guidance is given in Elliott 

(1977) for critical factors such as site selection, 

sample numbers, sampling design, and data anaiyse& To help investigators assess community 

impact, IJC (1988a) provides a partial list of 

literature descriptions of normal nearshore communities in habitats that most closely approximate 

Great Lakes AOCs. A detailed discussion of 

statistical methods is also included. 

Phase III of the detailed (Stage II) assessment 

consists of obtaining additional information concerning sediment toxicity (i.e., bioassays and fish 

histopathology) and stratigraphic characterization 

of sediment cores. A suite of bioassays is proposed for toxicological evaluation of sediments: 

Microtox - an acute, liquid-phase (eiutriate or pore-water) teat with luminescent 

bacteria (Bulich, 1984); 

Algal photosynthesis - an acute, liquidphase test using natural communities 

[algal fractionation bioassay (Munawar 

and Munawar, 1987)l or the laboratory 

species Selenasr?run cupricomutum (Ross 

et al., 1988); 

Zooplankton life-cycle tests (Duphniu 

magna liquid and solid phases) monitoring growth and reproduction (Nebeker et 

al., 1984; LBlanc and Surprenanf 1985); 

Chronic, solid-phase tests using the benthic invertebrates Chironomus tentons 

(Nebeker et ul., 1984), Hyulellu uztecu 

(Nebeker et al., 1984), or Wexugenio 

limbutu (Maiueg et al., 1983); 

A solid-phase fish bioaccumulation test 

with the fathead minnow Pimephalcs 

promelas (Mac et al., 1984) 

12-5 
Sediment Classi&ztion Methods Compendium 

m The liquid-phase (extract) Ames Suiinoneflalmicrosome assay, a bacterial mutagenicity test (Termant et al., 1987). 

In addition to bioassays, histopathological examinations of indigenous adult ftih (especially Ictolurus nebdosus), focusing on preneoplastic and 

neoplastic liver lesions (Couch and- Harshbarger, 

1985), are recommended. 

Also included in Phase Ill (Stage lI) work are 

chemical analyses and dating of sediment cores. 

Isotopic (“c 9b, UFe, l”(s) and biostratigraphic 

[i.e., ragweed (Ambrosia) pollen] methods are both 

recommended for dating sediment cores. This 

dating is necessary to establish the threedimensional configuration of the contaminated sediment 

mass and to assign a date to the sediment depositional unit. 

In Phase IV (Stage II) of the detailed assessment, studies on sediment dynamics are necessary 

to determine the following: 

U Potential water column impacts through 

resuspension; 

n Movement of contaminated sediment out 

of the AOC; 

n The quality and rate of new sediment 

accumulation; and 

w Vertical and horizontal redistribution of 

sediments and their contaminant burdens 

within an AOC 

This information is essential for the development 

and evaluation of a remediation plan. In the 

absence of practical predictive models, suspended 

sediment characterization (Poulton, 1987), shear 

strength measurements (Tenaghi and Peck, 1%7), 

and resuspension studies (Tsai and Lick, 1986) are 

recommended. 

12.2.1.2.3 Types of Data Required 

The Stage I initial assessment should be 

based on aberrant macrozoobenthic community 

structure (ascertained from family-level taxonomic identification); metals concentrations 

above background levels in the surficial sediments (ascertained from dating); hazardous persistent organic compound concentrations above 

detection levels in carp, benthos, or surficial 

sediments; metals concentrations in carp or 

benthos, established on a case-by-case basis; and 

presence in fishes of external abnormalities 

known to have contaminant-related etiologies. 

The Stage II detailed assessment should be 

based on a phased sampling of the physical, 

chemical, and biological aspects of the sediments The biological impacts should be assessed with both field (benthic invertebrate 

community structure and incidence of fish liver 

tumors) and laboratory (battery of selected 

bioassays) methods. The phased sampling 

approach will allow subsequent testing requirements to be reduced. When Phases I and II of 

Stage II have revealed homogeneous zones of 

sediment type and similar community structure, 

the number of Phase III samples can be appropriately scaled down. Impairment due to sediment contamination and the probable need for 

remediation are established when the biomonitoring results from the detailed assessment 

demonstrate significant departures from antrols. 

Each section of UC (1988a) contains a detailed discussion of the statistical procedures 

required, with references and examples. The 

preferred method of interpretation is left to the 

expert investigator in many cases. 

12.2.1.2.4 Necessary Hardware and Skills 

The initial assessment, and to an even 

greater degree the detailed assessment, requires 

a large array of field and laboratory equipment. 

Although none of the items recommended are 

unusual or inordinately sophisticated, one laboratory or field unit is unlikely to have all the 

required apparatus. Specific suggestions for 

hardware and skills are provided by DC (198&a). 

Because this approach is intended for major 

sediment assessment efforts, several groups 

would probably have to be mobilized to contribute to the effort. 

12-6 
12.2.1.3 Adequacy of Documentation 

Each component method described in UC 

(1988a) is fully referenced in the text and accompanied by a separate bibliography. Some methods 

are more developed than others, and areas where 

additional validation or calibration is needed are 

clearly identified in the text. 

12.2.2 Applkablllty of Method to Human 

Health, Aquntk Life, or Wlldllfe 

Protectloll 

The IJC strategy includes direct measures of 

effects on benthic infauna and fishes and is thus 

directly applicabie to aquatic biota. Existing 

sediment assessment methods (e.g., Apparent 

Effects Threshold, Sediment Quality Triad) could 

be used to evaluate the results of the Stage II 

detailed assessment and to determine whether 

chemically contaminated sediments have affected 

aquatic biota in the vicinity of AOCs. Although 

the UC (1988a) strategy was not designed to 

assess the effects of toxic chemicals on wildlife or 

humans, the tissue residue data and the sediment 

chemistry data may be useful in preliminary 

evaluations of contaminant exposure to these 

populations. Wildlife exposure could occur 

through consumption of chemically contaminated 

prey- Human exposure could occur through 

consumption of chemically contaminated fish or 

through dermal absorption by direct contact witb 

chemically contaminated sediments or water. 

12.23 Ability of Metbod to Generate 

Numerical Criteria for Specifk 

Chemicals 

The document was designed to provide guidance to assessment programs. Nevertheless, since 

chemical, toxicological, and infaunal data are 

collected in the Stage II assessment, it is possible 

that these data could be used to develop chemicalspecific criteria. For example, data from the Stage 

II assessment could be used to develop empirical 

sediment quality values (e.g., AETvalues) that are 

protective of aquatic biota in locations other than 

the AOC under consideration. 

123 USEFULNESS 

123.1 Environmental Applicability 

12.3.1.1 Suitability for Diflerent Sediment 

rrLpes 

The approach recommended in UC (1988a) is 

suitable for any sediment type. Indeed, one of its 

major objectives is to characterize and provide a 

three-dimensional map of the contaminated sadiment mass, including physical, chemical, and 

biological variables. Tbe investigator is given the 

flexibility to choose the appropriate sampling 

methods for the sediment type or types in &he 

AOC under study. 

12.3.1.2 Suitability for Different Chemicals 

or Classes of Chemicals 

The document is intended for situations where 

contamination is suspected, but where the toxic 

chemicals may or may not be identified. ‘I&e 

methods recommended by IJC (1988a) are effcctive for most contaminants found in Great Lakes 

sediments. ‘l%e broad-based nature of the ap 

proacb contains sufficient flexibility to deal with 

anomalous situations. 

12.3.1.3 Suitability for Predicting Eficts (RI 

Diierenr Organisms 

The proposed strategy includes both laboratq 

testing and analysis of indigenous communities 

(i.e., fsh, macrozoobenthos). In this way, laboratory results (i.e., chemistry, toxicity) that can be 

compared to standard anditions and literature 

values may be placed in the context of empirically 

derived effects data from the site under investigation. 

12.3.1.4 Suitability for In-Place Polhtanf 

Control 

The guidance document was developed specifically for the assessment of in-place pollutant 

problems. It is designed to fit into the framework 

of evaluating and dmosing remedial optioas by 

providing an adequate database on which to base 

12-7 
Sediment Chssification Methods Compendium 

such decisions. A companion document (UC, 

1988b) provides guidance in the selection of 

courses of remediation. 

12.3.1.5 Suitability for Source Control 

The detailed assessment provides an adequate 

framework for identifying hot spots, and for 

establishing significant differences from background conditions. In some cases, the resultant 

maps may provide further evidence of contaminant 

sources and migration patterns, using spatia! 

autocorrelation techniques. Presumably, such 

evidence could facilitate regulation of identified 

sources. However, source control is not a primary 

objective of the UC (1988a) strategy. 

12.3.1.6 Suitability for Disposuf Applications 

Although the document was not intended for 

the use in decision-making related to the disposal 

of material from navigational dredging, the data 

generated from an initial assessment could be used 

to make initial disposal decisions. Other practices 

for the assessment of dredged material may be 

more cost-effective, however. 

123.2 General Advantages and Umltations 

12.3.2.1 Ease of Use 

The proposed strategy is designed to be applicable to the AOC under investigation. It is intended 

to flexible, relying on the judgment and experience 

of those who apply it. A detailed assessment would 

be practical only in cases where a major remedial 

effort is contemplated. 

12.3.2.2 Relatiw Cost 

Y%e Stage I and II assesunents are costly 

compared to other less comprehensive methods of 

aeasing sediment quality. However, when compared to the potential remedial costs, the assessment 

a&s are relatively small. The sequential approach 

is designed to reduce sampling, analysis, and expense where possible. In many cases, the Stage I 

assessment need not be done. If it is clear that a 

sediment contamination problem exis& then the 

investigators may proceed diredly to Stage II 

assessmat. Alternatively, if the Stage I arxxssment 

produces no results of concern, then Stage Xl need 

notbeunderhken. Thecostofadetailedassessmea< although relatively high, is controlled somewhat by the sequential approach to data oolle&~~ 

No firm cost figures are atrrently available, but assessments planned for priority AOCs under Section 

118(c)(3) of the Water Quality Act of 1987 are 

projected to cost in the range of $500,000. T&se 

03sts are expeckd to vary from site to site. 

12.3.2.3 Tendency to Be Consd 

lbe strategy is designed to be highly pro&%ive 

of the environment. It combines chemical analysis, 

toxicity testing, and examination of indigenous 

communities to ensure that no significant effects are 

overlooked. Because the application of criteria is 

left to the expert judgment of the investigator, the 

degree of conservatism in decisioninaking will be 

variable. 

12.3.2.4 Level of Accepzance 

Ihe guidance doaunent (UC, 198&t) does not 

desuii a new method, but rather a combination of 

several types of methods, each widely accepted in 

its own sphere. The strategy as a whole is being 

usedforthefirsttimein 1989. 

12.3.2.5 Abilify to Be hnpk~nted by 

L4abora!ories with lypidw 

d Handling Fadties 

None of the methods is particularly unusual or 

dif.ficult, but the detailed assessmentrequhsa 

breadth of expert& and resourm that an individual 

organization may not possess. ‘Ibe strategy will 

needtobeimplementedbydtawingonavarietyof 

expert& in a given geographical area. 

12.3.2.6 Level of Effkt Required to Gkneme 

Re.SUit.9 

The total level of effort for a detailed assessment will be relatively high in most cases. ‘It& 

12-a 
strategy is most suitable for major evaluation 

projects. 

12.3.2.7 

Degree to which Results Lend Themselves to 

/nterpretarion 

The actual statistical gnalysis and interpretation to generate effects conclusions are relatively 

complex and should be done only by trained 

investigators. Specific statistical protocols are 

not recommended. However, the reader is given 

an array of choices, with comments on their 

respective strengths and weaknesses. The ultimate decision is left to the investigator. The 

inclusion of chemical, toxicological, and infauna1 information in the database allows the 

investigator to compare different types of indicators before making decisions. 

12.3.2.8 Degree of Environmental 

Applicability 

One of lhe strengths of a strategy that includes in situ community analysis is that effects 

data have a high degree of environmental relevance. Site-relevant species can even be substituted in the bioassay battery if necessary, and 

the body burden and community structure data 

are always site-specific. 

12.3.2.9 Degree of Accuracy and Precision 

The strategy proposed by the DC (1988a) is 

not a single method, but rather guidance for a 

study design containing many options and 

decision points. Overall precision or accuracy 

values would be impossible to calculate. Nevertheless, the criteria for selecting recommended 

protocols included a consideration of attainable 

precision. In many sections, the investigator is 

directed to choose the required level of precision 

for a given measurement during the study design 

process. The “accuracy” of an integrated strategy is difficult to assess, but the methods recommended by the IJC (1988a) were chosen for 

their relevance to the Great Lakes ecosystem, 

12.4 STATUS 

12.4.1 Extent of Use 

UC’s (1988a) document was published in 

December 1988 and distributed in early 1989. 

The strategy is intended for the Great Lakes, and 

was used for the first time in 1989. Most of the 

individual methods recommended are widely used 

and accepted. 

12.43 Extent to Wbkb the Approach Has 

Been Field-Validated 

The first extensive field validation of the approach was conducted in 1989-1991 as part of the 

ARCS program under section 118(c)(3) of the 

Water Quality Act of 1987. The ARCS Sediment 

assessment reports are expected to be released in 

1993. 

12.43 Reasons for Limited Use 

Most component protocols are in wide use. 

Because the UC (1988a) document describes a 

major effort with an integrated approach, the 

ARCS program is the only project where an 

undertaking using this approach has been initiated. 

12.4.4 Outlook for Future Use and 

Development 

With the backing of both signatories to the 

Great Lakes Water Quality Agreement, the document seems destined for widespread use in the 

Great Lakes basin. As methods progress, each 

section of the document will be updated. 

12.5 REFERENCES 

Adams, D.D., D.A. Darby, and R.J. Young. 1980. 

Selecfed analytical techniques for characterizing the metal chemistry and geology of fmegrained sediments and interstitial water. In: 

Contaminants and Sediments. RA. Baker (ed.) 

Ann Arbor Sci. Rub., Inc. Ann Arbor, MI. 

American Society for Testing and Materials. 

1%4. Procedures for testing soils. ASTM, 

12-9 
Stdiment Classificatbn Methods Compendium 

Philadelphia, PA. 535 pp. 

Bulich, A.A. 1984. Microtox - a bacterial toxicity test with general environmental applications. 

pp. 55-64. In: Toxicity Screening Procedures 

Using Baderial Systems. D. Xin and B.S. 

Dutka (eds.). Marcel Dekker, New York, NY. 

Cairns, J., Jr. 1986. The myth of the most sensitive species. Bioscience 36:670-672 

Chapman, P.M., and E.R. Long. 1983. The use 

of bioassays as part of a comprehensive approach to marine pollution assessment. Mar. 

Pollut. Bull. 14:81-84. 

Couch, IA, and J.C. Harshbarger. 1985. Effects 

of carcinogenic agents on aquatic animals: an 

environmental and experimental overview. 

Env. Carcinogen&s Rev. 3:63-105. 

Duncan, G-A., and G.G. LaHaie. 1979. Size 

analysis procedures used in the sedimentology 

laboratory, NWRI. Env. Can. NWRI contribution. 23 pp. 

Elliott, J.M. 1977. Some methods for the statistical analysis of samples of benthic invertebrates. Scientific Publication No. 25. Freshwater Biological Association. 160 pp. 

Green, R.H. 1979. Sampling design and statistical methods for environmental biologists. John 

Wiley and Sons, New York, NY. 257 pp. 

UC. 1988a. Procedures for the assessment of 

contaminated sediment problems in the Great 

Lakes. International Joint Commission, Windsor, Ontario, Canada. 140 pp. 

IJC. 1988b. Options for the remediation of contaminated sediments in the Great Lakes. 

International Joint Commission, Windsor, 

Ontario, Canada. 78 pp. 

LzBlanc, GA., and DJ. Surprenant. 1985. A 

method for assessing the toxicity of contaminated freshwater sediments. pp. 269-283. In: 

Aquatic Toxicology and Hazard Assessment, 

Seventh Symposium. R.D. Cardwell, R. Purdy, and R.C. Bahner (eds.), ASI’M STP 854. 

American Society for Testing and Materials, 

Philadelphia, PA. 

Mac, MJ., CC Edsall, RJ. Hesseiberg, and R.E. 

Sayers, Jr. 1984. Flow-through bioassay for 

measuring bioaccumulation of toxic substances 

from sediment. EPA DW-930095-01-O. U.S. 

Environmental Protection Agency, Chicago, IL. 

26 PP. 

Malueg, KW., G.S. Schuytema, J.H. Gakstatteq 

and D.F. ffiwczyk. 1983. Effect of Hexageniu on Lkphnia response in sediment toxicity tests. Env. Toxicol. Chem. 2:73-82. 

Munawar, M., and I.F. Munawar. 1987. Phytoplankton bioassays for evaluating toxicity of in 

sihc sediment contaminants. Hydrobiologia 

149~87-105. 

Nebeker, A.V., MA. Cairns, J.H. Gakstatter, KW. 

Malueg, and G.S.. Schuytema. 1984. Biological methods for determining toxicity of contaminated freshwater sediments to invertebra&s. Env. Toxicol. Chem. 33617630. 

Plumb, R.H., Jr. 1981. Procedures for handling 

and chemical analysis of sediment and water 

samples. Technical Report EPA/GE-81-1. 

U.S. Environmental Protection Agency/US. 

Amy Corps of Engineers Technical Committee 

on Criteria for Dredged and Fill Material, U.S. 

Army Waterways Experiment Station, Vicksburg, MS. 471 pp. 

Poulton, DJ. 1987. Trace contaminant status of 

Hamilton Harbor. J. Great Lakes Res. 13:193- 

201. 

Ross, PE., V. Jarry, and H. Sloterdijk. 1988. A 

rapid bioassay using the green alga Selenasbum capricotnutwn to screen for toxicity 

in St. Lawrence River sediments. Amefican 

Society for Testing and Materials. SI’P 

988:68-73. 

Schmitt, CJ., MA Ribick, J.L. Ludke, and T.W. 

May. 1983. National pesticide monitoring 

program: organochlorine residues in freshwater fish, 1976-79. Ftih and Wildlife Service 

Res. Publ. No. 152 U.S. Dept. of Interior, 

Washington, DC. 

Smith, S.B., MJ. Mac, A.E. MacCubbin, and J.C 

Harshbarger. 1988. External abnormalities 

and incidence of tumors in fish collected from 

three Great hkes Areas of Concern. Papcr 

presented at the 31st Conference on Great 

Lakes Research, McMastu University, Hamilton, Ontario. May 17-20,1988. 

Sokal, R.R., and F.J. Rohlf. 1969. Biometry. 

W.H. Freeman and Co., San Francisco, CA. 

Tennant, R.W., B.H. Margolin, D.D. SheIby, E. 

Zeiger, J.K. Haseman, J. Spalding, W. Caspary, 

12-10 
M. Resnick, S. Stasiewin, B. Anderson, and 

R. Minor. 1987. Prediction of chemical 

carcinogenicity in rodents from in situ genetic 

toxicity assays. Science 236:933-941. 

Terzaghi, K., and R.B. Peck. 1967. Soil mechanics in engineering practice. John Wiley and 

Sons, New York. 729 pp. 

Tsai, CH., and W. Lick. 1986. A portable 

device for measuring sediment resuspension. 

J. Great Lakes Res. 12:314-321. 

USEPA. 1984. Guidelines establishing test procedures for the analysis of pollutants under the 

Clean Water Act; final ruie and interim final 

rule and proposed rule. U.S. Environmental 

Protection Agency. Washington, DC. Federal 

Register Vol. 49, No. 209, Part VIII. pp. l210. 

Williams, J.D.H., H. Shear, and R.L. lttomas. 

1980. Availability to Scenedksmus quadricar& of different forms of phosphorus ia 

sedimentary materials in. the Great Lakes. 

Limnol. Oceanogr. 2!5:1-11. 

12-11 
CHAPTER 13 

Summary of Sediment-Testing Approach 

Used for Ocean Disposal 

David P. Redford 

U.S. Environmental Protection Agency 

499 South Capitol Street, SW (WH-556F), Washington, DC 20003 

(202)260-9179 

The Evaluation of Dredged Material Proposed for Ocean Disposal--Testing Manual 

(USEPA/USACE, 1991) commonly referred to 

as the “Green Book,” was published in February 

1991 by the U.S. Environmental Protection 

Agency (USEPA) and the U.S. Army Corps of 

Engineers (USACE). The Green Book contains 

national guidance for evaluating the suitability 

of dredged material for ocean disposal; it replaces the guidance of the original manual 

(USEPA/USACE, 1977) that was published by 

USEPA and the USACE in 1977. The manual 

stresses the use of bioassay and bioaccumulation 

testing as evaluative tools, and it contains technical guidance on the use of such tests. The 

following is a summary of the 1991 manual and 

the approach used by USEPA and the USACE 

to determine the suitability of dredged material 

for ocean disposal. The manual will be revised 

at a future date, based on the findings of an 

EPA Science Advisory Board (SAB) review 

(SAB,1992), and changes will be made to the 

Ocean Dumping Regulations (referenced below). 

13.1 APPLICATION 

The 1991 USEPA/USACE Green Book 

provides updated guidance for dredging applicants, scientists, and regulators to evaluate 

dredge-material compliance with the 1977 U.S. 

Ocean Dumping Regulations [Title 40, Code of 

Federal Regulations (CFR), Parts 220-228]. The 

manual is applicable to all activities involving 

the transportation of dredged material for the 

purpose of dumping it in ocean waters outside 

the baseline from which the territorial sea is 

measured. The guidance in this manual is applicable to dredging operations conducted under 

permits as well as to federal projects conducted 

by the USACE. The procedures in this manual 

do not apply to activities excluded by 

40 CFR 220.1. 

It is important to note that the regulations 

are legally binding and that the guidance provided in this manual is responsive to the specific 

requirements of these regulations, but the manual does not carry the force of law. The document simply provides guidance on evaluating the 

potential environmental impact of dredged-material ocean disposal. 

The manual is organized into tiers for efficient evaluation of the suitability of dredged 

material for ocean disposal. Within the tiers, 

specific physical, chemical, and biological tests 

are recommended. To meet specific regional 

needs, USEPA Region and USACE District 

offices are to develop local agreements and 

manuals to implement the national guidance in 

the 1991 Green Book (such as using local species in biological tests and screening for particular contaminants in chemical analyses). 

13.1.1 current use 

The 1991 Green Book replaces the 1977 

Green Book. USEPA Region and USACE District offices are developing local agreements and 

regional testing manuals that implement the 

1991 Green Book guidance and establish permit 

procedures for dredging and dredged-material 

disposal. 

Projects that have been issued under 

USACE permits prior to the completion of the 

new local agreement/manual for the area covered by the project may continue to be evaluated 
Sediment Classification Methods Compendium 

according to the 1977 guidance manual and the 

existing local guidance. New dredged-material 

disposal projects, projects that have not had 

sampling and analysis plans approved prior to 

finalization of the local agreement/manual, 

should be evaluated under the updated guidance 

in the 1991 Green Book. Ongoing projects that 

have been approved based on 1977 Green Book 

guidance should be reevaluated according to 

1991 Green Book guidance and the new local 

agreement/manual within 3 years of permit 

approval. 

13.1.2 Potential Use 

The Green Book guidance, and revisions 

thereof, will be applied to dredged-material 

evaluations for the foreseeable future. 

The manual will be revised at a future date 

based on (1) the findings of an EPA SAB review (SAB, 1992), (2) technical advances in 

assessing sediment contamination and marine 

environmental impact, and (3) changes to the 

Ocean Dumping Regulations. 

13.2 DESCRIPTION 

Analysis of sediment to determine its suitability for ocean disposal is conducted according to the procedures in the 1991 Green Book. 

The 1991 Green Book recommends procedures 

that satisfy section 103 of the Marine Protection, Research, and Sanctuaries Act of 1972 

(MPRSA), Public Law 92-532. The MPRSA 

was enacted to regulate ocean dumping of all 

materials that might adversely affect human 

health, the marine environment, or other legitimate uses of the oceans. In addition, the 

MPRSA implements the Convention on the 

Prevention of Marine Pollution by Dumping of 

Wastes and Other Matter (London Dumping 

Convention), of which the United States is a 

signatory. MPRSA section 103 specifies that 

all proposed operations involving the transportation and dumping of dredged material 

into ocean waters must be evaluated to determine the potential environmental impact of 

such activities. These environmental evaluations must be in agreement with the criteria 

published in 40 CPR Parts 220-228 and 

33 CFR Parts 320-330 and 335-338. 

Technical guidance on specific methods 

for testing dredged material is presented in the 

1991 Green Book. If the results of the appropriate tests show that the proposed dredged 

material meets the chemical- and biologicaleffects criteria, and meets other requirements 

in the regulations, disposal of the material at a 

designated ocean dredged-material disposal site 

(ODMDS) is supported. If the test results 

show that the material does not meet the criteria set forth in the regulations, significant 

impact on the ocean environment is predicted. 

Significant adverse impact may include adverse consequences to the marine ecosystem 

and negative human-health effects from uses 

of the marine environment. 

The manual does not present guidance for 

the disposal of dredged material that fails to 

meet the regulatory criteria. Such disposal 

involves management decisions and case-specific engineering work (e.g., control of dump 

releases, disposal-site capping, submarine 

burial, and predisposal treatment) that are 

beyond the scope of the document. 

13.2.1 Description of Method 

Integral to the 1991 Green Book is a 

tiered-testing procedure to characterize 

dredged material and predict its impact on the 

water-column and benthic environment at 

ODMDSs. The procedure was developed by 

USEPA and USACE personnel and testinglaboratory researchers, and is consistent with 

the requirements of the Ocean Dumping Regulations, state-of-the-art dredged-material 

evaluation techniques, and the realities of the 

testing and permitting process for new and 

existing projects. Knowledge of local conditions is both recommended and necessary to 

adapt the national guidance in the manual to 

specific dredged-material projects. USEPA 

Regions and USACE Districts are presently 

developing local agreements/manuals to apply 

13-2 
13-“Green Book” Sediment-Testing Approach 

the national guidance of the manual to specific 

dredging and disposal areas. 

The tiered-testing procedure in the Green 

Book comprises four tiers, with decision points 

at each tier (Figure 13-1). Each successive tier 

provides increasing investigative intensity to 

generate the information for permitting decisions on ocean disposal. 

The tiered-testing procedure is constructed 

to determine whether the dredged material 

meets the limiting permissible concentration 

(LPC), as defined in section 227.27 of the 

Ocean Dumping Regulations. The LPC for the 

liquid-phase concentration of dredged material 

in the water column is the concentration that, 

after allowance for initial mixing, does not exceed applicable marine water-quality criteria 

(WQC) or a toxicity lhreshald of 0.01 of the 

acutely toxic concentration. The LPC of the 

suspended particulate and solid phases is the 

concentration that will not cause unreasonable 

toxicity or bioaccumulation. 

The overall tiered-testing procedure is relatively flexible. The dredged-material evaluator 

can enter and exit the testing procedures at any 

tier. However, to begin the evaluation in Tier 

II, III, or IV, the existing data must satisfy the 

requirements of the earlier tier(s). Additionally, Tier II testing for water-quality criteria 

(WQC) compliance is mandatory if the watercolumn evaluation cannot be completed within 

Tier I. To exit any tier before reaching a 

decision on LPC compliance, the dredgedmaterial evaluator must select an option other 

than open-ocean disposal. 

In most cases, determinations of LPC compliance can be made in Tier I, II, or III. In 

extraordinary cases, where LPC compliance 

cannot be determined by Tier III, the dredged 

material must be evaluated under Tier IV. 

Tier IV tests are case-specific investigations of 

potential impact of the dredged material at the 

ODMDS. Significant investment in the research and development of analytical methods 

is usually necessary to conduct Tier IV evaluations, and the applicant might select an alternative to open-ocean disposal instead of 

proceeding with Tier IV testing. Similarly, an 

applicant can try to save time and money by 

proceeding directly to Tier II, III, or IV if it is 

believed that analysis in the earlier tiers will 

not lead to a definitive evaluation. The only 

absolute requirement is that the dredged material must comply with the regulations if it is to 

be dumped at an ODMDS. The tiered-testing 

procedure facilitates this determination. 

In summary, the 1991 Green Book 

Includes state-of-the-art methods to 

determine the potential impact of marine-sediment disposal; 

Ensures adherence to the Ocean Dumping Regulations (40 CFR Parts 220- 

228); 

Incorporates existing (and valuable) 

regional expertise and guidance into 

the evaluation process; and 

Provides for National consistency in 

evaluating dredged material for ocean 

disposal. 

13.2.1.Z Objectives and Assumptions 

The objective of the tiered-testing procedure is to determine whether the water-column 

and benthic LPC is met for the proposed 

dredged material, as defiied in the Ocean 

Dumping Regulations. Three decision options 

are possible as the dredged-material evaluator 

proceeds through the tiers. 

(1) The LPC is met; the ocean disposal 

option is supported; further evaluation 

is unnecessary. 

(2) The LPC evaluation is inconclusive; 

the ocean disposal option is not supported; proceed to the next tier. 

(3) The LPC is not met; the ocean disposal 

option is not supported; further evaluation is unnecessary. 

133 
Sdimd Class@dion Methods Compendium 

KfYTOWYEWCUfLRC 

I' 

- 

- 

-- 

TIER I 

TKRS LE.LP 

igun 1Sl. 1991 Green Book tlered-testlng procedure. 
13--“Green Book” Sediment-Testing Appvach 

Both the water-column and bentbic LPC 

considerations must be satisfactorily resolved for 

the open-ocean disposal option to be supported. 

An inconclusive evaluation in Tiers I-III requires 

the dredging applicant to conduct additional 

testing in subsequent tiers, or to decide not to 

oceandump. However, a determination of LPC 

noncompliance does not necessarily exclude all 

possibilities for ocean disposal. Management 

actions might be feasible to make the dredged 

material meet the LPC. Management actions for 

dredged material that exceeds water-column or 

bentbic LPC are rrot included in the Green Book 

because of the wide range of available options 

and the project-specific nature of such work. 

It is assumed that the users of the 1991 

Green Book are generally familiar with the need 

for and methods of dredged-material testing. 

The manual is not a standalone document. The 

guidance in the manual requires the evaluator to 

consult the regulations frequently (40 CFR Parts 

220-228 is included in the Green Book as Appendix A) and to have a general understanding 

of material contained in the numerous citations 

and references. The guidance in the manual 

concentrates on data collection and decision 

points, and it only summarizes recommended 

field and laboratory procedures that can be used 

to obtain data. The user must refer to the original sources for most of the physical, chemical, 

and biological testing procedures. 

13.2.1.2 Level ofEflorr 

Tier I: Initial Assessments-Tier I is used to 

identify contaminants of concern and determine 

dredged-material LPC compliance through analysis of existing physical, chemical, and biological information. For many dredging projects, 

there is a wealth of readily available information 

on the proposed dredged material and on the 

characteristics of the disposal site. This is especially true of areas that have historically undergone maintenance dredging or have been the 

subject of other studies, such as fishery assessments. The available information for a given 

area might not be sufficient to reach a final LPC 

evaluation, but often there are accessible highquality data that can supplement the results of 

tests in subsequent tiers and facilitate reaching 

an early decision with lowered expenditure of 

time and resources. 

Whatever the source of information for 

Tier I evaluations, the quality of the data must 

be evaluated and weighed accordingly. The 

references in Chapter 13 of the manual, QualityAssurance Considerations, should be consulted 

for guidance for evaluating the quality of data 

ob&ned from different information sources. 

If the information set compiled in Tier I is 

complete and comparable to information that 

would appropriately satisfy the LPC in Tier II, 

III, or IV, a decision on regulatory compliance 

be completed without proceeding into the next 

tiers. For compliance determination to be completed within Tier I, the weight of evidence of 

the collected information must convincingly 

show that the dredged-material disposal either 

will or will not meet tbe LPC. 

Included in Tier I is an assessment of the 

three exclusionary criteria in 40 CFR 227.13(b): 

(1) the dredged material is predominantly sand, 

gravel, or rock from a highenergy area; (2) the 

material is suitable for beach nourishment; or 

(3) the material is similar to the disposal site 

and from an area far removed from pollution 

sources. If one or more of the above exclusionary criteria can be satisfied, the LPC is met for 

the dredged material and no further evaluation is 

required. If none of the exclusionary criteria is 

met and the collected information is insufficient 

to reach a definitive LPC determination, the 

evaluation process moves to Tier II. 

Tier II: PbysScaUCbemical EvaluationcTier 

II consists of physical and chemical data evalua- 

.tion. To determine marine WQC compliance, a 

numerical mixing model is used; to evaluate 

benthic-impact potential for nonpolar organic 

compounds, a theoretical bioaccumuIation potential (TBP) calculation is used. The conceptual 

purpose of the tier is to provide reliable, rapid 

screening of impact potential without the need 

for further testing. This purpose is fulfilled for 

water-column evaluations, but at present there is 

no USEPA-approved single screening procedure 
Sediment ClassiJication Methods Compendium 

for deposited sediment. When technically sound 

sediment-quality criteria (SQC) are developed 

and approved for dredged-material evaluation, 

they will be incorporated at this level. 

Tier II: Water-Column PbysicaVCbemicaI 

Evaluation=The Tier II water-column evaluation for WQC compliance is a two-step process that includes the application of a numerical 

mixing model. in Step 1, the model is used as a 

screen; all of the contaminants in the dredged 

material are assumed to be released into the 

water column during the disposal process. If the 

model predicts that the concentration of contaminants of concern released into the water column 

is less than the applicable WQC and if no synergistic effects among the contaminants are 

suspected, the dredged material meets the watercolumn LPC and no further water-column evaluations are necessary. 

If LPC compliance cannot be shown in 

Step 1, Step 2 is conducted. In Step 2, chemical 

data from an elutriate test of the dredged material are run in the model. Compared to the assumption of total contaminant release in the Step 

1 screen, the elutriate data applied in Step 2 are 

a more precise representation of the concentration of contaminants that would actually be 

released into the water column during ocean disposal of dredged material. 

If the model predicts in Step 2 that any 

WQC are exceeded, the water-column LPC is 

not met (open-ocean disposal not supported). If 

there are WQC for all of the contaminants of 

concern, if no WQC are exceeded by the Step 2 

model, and if no contaminant synergistic effects 

are suspected, the water-column LPC is met and 

no further water-column evaluations are necessary (open-ocean disposal supported). If there 

are contaminants of concern without WQC or if 

synergistic effects are suspected, water-column 

toxicity and water-column LPC compliance must 

be evaluated in Tier III. 

Numerid Models for Initial Mixtng-Numerical models are used to evaluate dredged-material dilution during the initial-mixing phase of 

ocean disposal, as defined in the regulations. 

The 1991 Green Book recommends using the 

USACE Automated Dredging and Disposal 

Alternatives Management System (ADDAMS) 

models to evaluate initial mixing of dredged 

material at ODMDSs. ADDAMS models can be 

run on a personal computer with a minimum of 

hardware. The models account for the physical 

processes of dredged-material disposal at openwater disposal sites by calculating the watercolumn concentrations of dissolved contaminants 

and suspended sediments and the initial deposition of material ‘on the bottom. Three separate 

ADDAMS models address different methods of 

disposal: 

n DIFID Disposal from an instantaneous dump 

n DIFCD Disposal from a continuous 

discharge 

n DIFHD Disposal from a hopper 

dredge 

To evaluate initial mixing following ocean 

disposal, the appropriate model is run for &e 

contaminant requiring the greatest amount of 

dilution to meet the LPC. The models simulate 

movement of the disposed material as it falls 

through the water column, as it is transported 

and diffused by the ambient current, and as it 

spreads over the bottom. The models have 

some limitations; for example, the DIFID model 

will not work for very shallow disposal sites 

where the discharge time from the barge exceeds 

the descent period to the bottom. However, the 

models can simulate a wide range of disposal 

options. USEPA and the USACE are in the 

process of field-verifying these models. 

Appendix B of the 1991 Green Book is a 

summary of the ADDAMS models; the armputer diskettes that accompany the manual anttaln 

the models themselves. ADDAMS modeling 

personnel at the USACE Waterways Experiment 

Station (WES), Vicksburg, Mississippi, are 

available to supply model updates, answer questions, and assist with the selection and running 

of the individual models. 
13-“Green Book” Se&nent-Testing Approach 

Table 13-l. 1991 Green Book Species for Water-Column and Benthic Evaluations. 

- 

Water-Column Species Benthlc Species 

w Crustaceans 

Mysids 

Mysidopsis sp.’ 

Neomysis sp.’ 

ffolmesimysis sp.’ 

Shrimp 

Palaemonetes sp. 

Penaeus sp. 

Panda/us sp. 

Crab 

Caliinectes sapidus 

Cancer sp. 

n Fish 

Menidia sp .’ 

Cyma togaster aggregata’ 

Cyprinodon variegatus 

bgodon rhomboides 

Leiostomus xanthurus 

Citharicthys stigmaeus 

Leuresthes tenuis 

Coryphaena hippurus 

n Zooplankton 

Copepods 

Acartia sp.’ 

Mussel larvae 

Mytilus eduli9 

Oyster larvae 

Cfassostrea virginit 

Osfrea sp.’ 

Sea-urchin larvae 

Strongyiocentrotus 

purpufa tus 

Lytechinus pictus 

n Crustaceans 

lnfaunal Amphipods 

Rhepoxynius sp. 

Ampeliscf3 sp. 

Eohaustorius sp: 

Grandiderella japonica 

Corophium insidiosum 

Mysids 

h4ysidopsis sp. 

Neomysis sp . 

Holmesimysis sp . 

Shrimp 

Penaeus sp. 

Paiaemonetes sp. 

Crangon sp. 

Panda/us sp. 

+yonia ingentis 

Crab 

Calinectes sapidus 

Cancer sp. 

8 Fwh 

Clevela*ia ios 

Atherinops affinis 

= Burrowing Polychaetes 

Neanthes sp: 

Nereis sp.’ 

Nephthys sp. 

G&era sp. 

Ar8nicok sp. 

Abafenico/a sp. 

n Moulws 

Yokiia limatula 

Macoma sp. 

Nucula sp. 

Profothaca staminea 

Tepes japonica 

Uercenariti mercenaria 

%commended test species. 

13-7 
Sediment Chssificu tion Methods Compendium 

The model outpul can present water-column 

contaminant concentrations in milligrams per 

liter. These concentrations are compared to the 

appropriate LPCs to determine compliance. 

Tier II: Benthic PhysicaUChemical Evaluations-As previously noted, only benthic effects 

attributi to nonpolar organic chemicals in the 

deposited sediment can be addressed in Tier II at 

the present time. Nonpolar organic chemicals include all organic compounds that do not dissociate 

or form ions. T&se include chlorinated hydrocarbon pesticides, other halogenated hydrocarbons, 

polychlorinated biphenyls (PCBs), most polynuclear aromatic hydrocarbons (PAHs), dioxins, and 

furans. It does not incIude polar organic compounds, organometals, and metals. If all of the 

contaminants of ancem in the dredged material 

are nonpolar organic compounds, the lheoretical 

bioaccumulation potential (TBP) can be calculated 

for the dredged material and the reference sediment* to determine benlhic LPC compliance ne 

TBP calculalion is an environmentally conservative screen, based on calculating the concentration 

of the nonpolar organic chemical in the sedimenl, 

the total organic&on concentration, and the 

percent lipid content of an organism of interest. If 

the TBP of the dredged material is not staGstically 

greater than that of the reference material, the 

LPC for the nonpolar organic contaminants is met. 

(Aa&-toxicity evalualions must be performed 

under Tier Ill unless sufficient toxicity information 

was obtained under Tier I.) 

If any of the contaminants of concern are 

polar organic compounds or have suspeded toxic 

components or if the dredged-material TBP exceeds the reference-material TBP described above, 

the bioacarmulation evaluation for benthic impact 

by the dredged material must take place in Tier III 

or IV. ‘The benefit of additional tests in Tier II to 

screen for benthic impact is recognized by USEPA 

IA reference sediment is detined as a sediment, wbstantially free of ax’itaminank, that is as similar as practicable to 

the grain size of the dredged material and the sediment at 

the disposal site, and that reflects the conditions that would 

exist in tbe vicinity of the disposal site had no dredgedmaterial disposal ever taken place, but bad all other influ-. 

enceS on sediment condition taken place. 

and the USACE, and new tests are under develop 

ment and evaluation. When the scientific and 

regulatory community verifies one or more of 

these tests, they will be incorporated intoTier II in 

a future Green Book revision. Meanwhile, evaluation of benthic impad that cannot be ma& in 

Tier I must be completed in Tier III or IV. 

‘Ikr III: Biological Evduatioas-73er III tests 

include (1) determination of w&r-column toxicity 

and (2) assessment of contaminant toxicity and 

bioaccumulation from the material to be dredged. 

The evaluations in this tier are based on the output 

from Tiers I and II and comprise standardized 

bioassays wilh the organisms listed in Table 13-1. 

Tier III: Water-Column Biological Evaluations-Tier III water-column tests are acute tests 

that evaluate the toxicity of the dissolved and 

suspended portions of the dredged material that 

remains in the water column after initial mixing. 

The bioassays are run if the Tier II evaluations 

are inconclusive, e.g., if there are not applicable 

WQC for all contaminants of concern or there .is 

reason to suspect synergistic effects among the 

contaminants. (See Tier II.) The tests involve 

exposing fish, crustaceans, and zooplankton to a 

dilution series containing both dissolved- and 

suspended-sediment components of tie dredged 

material. A typical test monitors organism mortality over a 96-h period. 

The results of the bioassays are used to calculate the &, concentration of the dredged mated 

in the water c&unn. The LPC for this evaluation 

is 1 percent of the L&, outside the ODMDS 

during the initial 4-h mixing period and anywhere 

in the marine environment 4 b after disposal. 

Following Ihe determination of the LPC for the 

proposed dredged material, the data are used to 

ND ihe numerical model (see model discussion 

above) and determine LPC compliance. 

‘Ikr III: Benthic Biological Evaluations-Benthic evaluations in Tier III consist of toxicity and 

bioaccumulation tests. To amdud these tests, the 

1991 Green Book provides laboratory guidance 

on sediment preparation; treatment, reference-, 

and control-sediment tests; ,replicates; organism 

13-8 
It “Green Book ” Sediment-Testing Approach 

handling; test-chamber conditions; QA considerations; and data analysis. The organisms used in 

the tests are surrogates for disposal-site species 

and are used to estimate dredged-material effects. 

The toxicity tests quantify mortality. If the mortality of the test species in the dredged-material 

bioassays is greater than the allowable percentage 

over the mortality in the reference-sediment bioassays, the LPC is not met. If, however, the 

dredged-material tests below the allowable perantage, or the increased mortality is statistically 

insignificanf the LPC is met. 

The bioaccumulation tests evaluate the potential of benthic organisms to accumulate contaminants from the dredged material in their tissues. 

At the conclusion of the tests, the tissues of the 

organisms are analyzed for the contaminants of 

concern that are identified in Tier I. 

Section 227.27 of the Ocean Dumping Regulations requires that bentbic bioassays be conducted on dredged material with filter-feeding, 

deposit-feeding, and burrowing species. Infaunal 

amphipods, such as Ampelisca sp. and Rhepoxyn&r sp., are sensitive bioindicators and strongly 

recommended in the Green Book as the preferred 

species for toxicity tests. Infaunal amphipods 

filter-feed, deposit-feed, and, to some extent, burrow in the sediment, thereby fulfilling the three 

organism otegories in the regulations. For bioaccumulation evaluations, the manual recommends 

using a burrowing polychaete (e.g., Neanrhes sp. 

or Nereis sp.) and a deposit-feeding bivalve moliusc (e.g., Macoma sp. or Yoldia limaMu). In 

summary, the manual recommends that at least 

two species be tested for acute toxicity and at 

least two other species for bioaccumulation evaluation. Each set of test species should cover the 

three species types stipulated in the regulations. 

The ecological and economic relevance of the 

organisms and the practical aspects of using the 

species in the laboratory, such as tolerance to 

grairi-size ranges and seasonal availability, also 

must be considered when selecting the test 

species. 

TIte Tier III bioaccumulation evaluation compares the contaminant level in the tissues of the 

organisms to two criteria: (1) the United States 

Food and Drug Administration (FDA) Action 

Levels for Poisonous or Deleterious Substances in 

Fish and Shellfish for Human Consumption and 

(2) the contaminant levels in organisms that are 

exposed to the reference sediment. Regardless of 

the statistical comparison to the reference-material 

test organisms, if the level in the tissues of 

dredged-material organisms statistically exceeds 

$e FDA levels in any’ category, the LPC is not 

met. If the dredged-material results are lower 

than the FDA action levels and not statistically 

greater than the reference material level, the LPC 

for bioaccumulation is satisfied. However, if bioaccumulation exceeds that found in the referencematerial tests, the test results must be evaluated 

against case-specific criteria. USEPA and the 

USACE develop the evaluative criteria case by 

case from local technical information that addresses the bioaccumulation aspects of the benthic 

criteria of section 227.13(c)(3) of the regulations. 

At present, tests for chronic sublethal exposure to benthic contaminants are being developed. 

When the tests are approved by USEPA, they 

will be incorporated in Tier III in future updates 

to the Green Book. 

Tier IV: Advanced Biological EvalurtionsTier IV consists of bioassay and bioaccumulation 

tests to evaluate the long-term benthic arid watercolumn impact of dredged material. Tests at this 

level are selected to address specific issues for a 

specific dredging operation that could not be fully 

evaluated in the earlier tiers. Since these tests are 

case-specific and since they require significant 

time and money to complete, evaluative criteria 

must be agreed on in advance by USEPA and by 

the USACE to determine compliance with the 

UK. 

Conducting Tier IV benthic testing is possible 

with current methods, but the 1991 Green Book 

emphasizes that this tier is not intended for routine application. Tier IV benthic tests consume 

significant resources of the dredging applicant 

and of the regulatory authority, and a fiial noncompliance determination is still possible. l’herefore, the applicant must weigh the options and 

decide whether to perform Tier IV testing or to 

consider an alternative that does not involve 

ocean dumping, such as upland disposal. If the 

13-9 
Sediment Classifiultion Methods Compendium 

applicant elects to proceed with Tier IV testing, 

the role of the regulatory authority is to design 

tests that lead to a defmitive.LPC evaluation for 

the project. 

Under Tier IV evaluations, bioaccumulation 

testing measures the steady-state body burden of 

contaminants of concern in tissues of organisms 

subjected to long-term laboratory exposures or in 

tissues of appropriately sampled field organisms. 

Tbe contaminant concentration in tbe tissues of 

dredged-material test organisms is compared 

against the appropriate FDA action levels and 

against bioaccumulation data obtained from organisms that are exposed to reference-material 

sediment. If contaminant bioaccumulation in the 

dredged-material organisms is less than the FDA 

levels but greater than the levels in the referencematerial organisms, organisms are collected from 

the vicinity of the disposal site and analyzed for 

the contaminants of concern. If the contaminant 

bioaccumulation of the dredged-material organisms is lower than the steady-state body burden 

of the fieldcollected organisms, the LPC for bioaccumulation is met. If field-collected organisms 

have contaminant levels lower than those of the 

dredged-material organisms, case-specific criteria 

are developed to make a final LPC compliance 

determination for bioaccumulation. 

13.2.1.2.1 Type of Sampling Required 

Section 8.0 of the 1991 Green Book, Collection and Preservation of Samples, provides general information on sampling plans and sample 

handling, preservation, and storage. 

To adequately and efficiently conduct a 

dredged-material evaluation, a comprehensive 

sampling plan should be in place before sampling 

begins. Sufficient amounts of sediment and water 

should be collected to conduct the necessary evaluations. Carehrl consideration of maximum 

allowable and recommended holding times for 

sediments, as well as the exigencies of resamp 

ling, should.be given careful consideration. Additionally, sample size should be small enough to 

be ccrnveniently handled and transported, but 

large enough to meet the requirements for all 

planned analyses. Tbe overall confidence of the 

final LPC determination is based on the following 

three factors. 

n Collecting representative samples; 

m Using appropriate sampling tedmiques; 

and 

n Protecting or preserving the samples until 

they are tested. 

Table 13-2 shows tbe general sampling requirements to conduct dredged-material testing. 

Actual sampling requirements are projectspecific 

and are determined during the development of the 

project plan, based on the guidance that is provided in the 1991 Green Book and in local agreements/manuals. 

13.2.1.2.2 Methods 

As described in Section 13.2.1.2.1 above, 

only existing information is evaluated in Tier I, 

This requires the careful compilation and analysis 

of such information. If the information czumot 

show that the proposed dredged material meets 

one of the exclusionary criteria, or if the 

information is insufficient to reach an LPC determination, physical, chemical, and biological information on the dredged material and the ODMDS 

must be collected in Tiers II and/or III. 

Proper sample collection, handling, and preservation are critical to the accurate evaluation of 

Tier II and III test results. Sampling methods are 

usually developed by individual testing laborab 

ries and documented in standard operating procedure (SOP) documents. Consistent use of SOPS 

in the field and laboratory ensure that sampling 

and analytical errors are minimized. 

Methods necessary to conduct toxicity and 

bioaccumulation evaluations may include the 

following: 

n Sieving; 

I Combustion; 

n Gravimetry; 

13-10 
Xi-“Green Book” Scdimpnf-Testing Approach 

Table lS2. Sample-Collection Requirements 

ests Water Samples Sediment Samples 

Disposal Dredging Dredging Reference 

Site Site COtltd’ Site Site control’ 

ler II 

Water Column 

Screen 0 0 

Elutrlate 0 0 Cl 

ler II 

Benthic 0 a 

ler 111 

Water Column q b cl 0 0 

ler 111 

Benthic 0 0 q 

ler IV 

Water Column Cl 0 0 0 

ler IV 

Benthic 0 0 a 

01 

‘May or may not have to be tield-collected. 

bDilution water; disposal-site water, artificial water, or clean seawater. 

Gas chromatography (GC); 

Electron-capture detedion (ECD); 

Mass spectrometry (MS); 

Graphite furnace atomic absorption spectroscopy (GFM); 

Atomic absorption spectroscopy (AAS); 

Inductively coupled plasma (ICP) technique; 

%-h elutriate toxicity bioassays; 

lo-day whole-sediment toxicity bio- 

=ys; 

n lOday whole-sediment bioaccllmulation 

tests (for Uace-metals analysis only); and 

n Z&day whole-sediment bioaccumulation 

tests. 

Project-specific methods necessary to 

conduct Tier IV water-column and benthic 

evaluations may include laboratory and/or field 

evaluations of long-term toxicity or bioaccumulation effects of the dredged material, 

such as the following: 

n Population-survival assessments; 

m Community-change assessments; and 

l Reproduction assessments. 
Sediment Clnss$c~~tion Methods Compendium 

1321.23 Types of Data Required 

As discus& in Sections 13.2.12.1-13.2.1.2.4 

above, data required to conduct the WC evaluations may include the following: 

Physical sediment data; 

Organic- and inorganic-chemistry 

ment data; 

Organic- and inorganic-chemistry 

ment-elutriate data; 

Physical-oceanography data; 

Bioassay data; 

Bioaccumulation data; and 

Field species data. 

13.2.1.2.4 Necessary Hardware and Skills 

sedisediThe hardware and skills necessary to amdud 

1991 Green Book evaluations are relatively specialized. Many federal, state, and contract laboratories have capabilities to conduct most or all of 

the necessary evaluations. However, to conserve 

time and resources, field sampling, laboratory 

work, data management, and analysis of the results are often conducted by separate organizations 

according to aptitude, cost, and scheduling 

parameters. 

The general categories of capabilities necessary to reach a Tier III dredged-material IJC 

compliance determination are the following: 

= Regulation and literature research; 

n Field sampling at the dredging site, disposal site, and reference site; 

W Physical analysis of sediment samples; 

n Trace-metal (chemical) analysis of water 

and sediment samples; 

Organic-compound (chemical) analysis of 

water and sediment samples; 

Numerical modeling for initial-mixing 

WlySk, 

Toxicity bioassay testing of elutriate samples; 

Toxicity bioassay testing of whole-sediment samples; 

Bioaccumulation testing; 

Chemical analysis of tissue samples; 

Statistical analysis of test results; 

Quality-assurance implementation 

(throughout evaluation); and 

Compliance determination. 

13.2.1.3 Dtxxune~tion 

Throughout the 1991 Green Book references 

are provided for the recomm ended sampling and 

testing methods, data analyses, QA procedures, 

and additional testing guidance. For convenience 

to manual usefs, a copy of the U.S. Qcean Dump 

ing Regulations (40 CFR Parts 220-228) is includedinthe1991GreenBookasAppettdixA 

Information on documentation and recordkeeping is interspersed throughout the testing 

guidance. Records ensure that all aspects of the 

field and laboratory work are documented so that 

the resulting data may be properly interpreted 

Dredged-material test data may be rejected if their 

history cannot be confidently traced. 

‘lk2.2 Applicability of Metbod to Homae 

Health, Marine Life, or Wildlife 

PI-Ot4XtiOB 

The effects-based guidance provided in the 

1991 Green Book is directly applicable to the 

protection of human health, marine life, and 

13-12 
13-"Green Book” Sediment-Testing Approach 

wildlife because it is based on determining LPC 

compliance. if the testing shows that either the 

UC for the water-column or benthic environment will be exceeded, ocean disposal for the 

proposed dredged material is not supported. In 

40 CFR 227.27(a), the LPC is defined as the 

concentration of the liquid phase of the dredged 

material that will not exceed either the established WQC or 1 percent of the acutely toxic 

concentration following the initial-mixing phase 

(initial mixing is defined in 40 CFR 229.29). In 

40 CFR 227.27(b), the LPCS for the suspended 

particulate and solid phases are defined as those 

concentrations “. . . that will not cause unreasonable acute or chronic toxicity or other sublethal adverse effects based on bioassay results 

using appropriately sensitive marine organisms 

. . . or will not cause accumulation of taric 

materials in the human food chain.” 

The tiered-testing procedure in the manual 

establishes a conservative, yet workable, decision-making process for environmentally protective dredged-material management. Dredged 

material that poses no risk of adverse impact is 

readily supported for ocean disposal early in the 

procedure (i.e., Tier I or 11). Dredged material 

that has unknown impact potential is evaluated 

to the level required to make a definitive LPC 

compliance determination. Only dredged material that is shown to meet both the water-column 

and benthic LPC through state-of-the-art analytical techniques is supported for open-ocean 

disposal. 

13.23 Ability of the Testing to Generate 

Numerical Criteria for Specific 

Chemicals 

The physical, chemical, and biological data 

generated by the Tier II, III, and IV tests can be 

used lo field-validate SQC that are presently 

under development. The state-of-the-art sampling and analytical techniques contained in the 

1991 Green Book guidance wiI1 provide for increases in method reproducibility, confidence of 

the test data, and utility to SQC research and 

development projects. 

133 USErmLNESS 

133.1 Environmentd AppUabiHty 

The guidance in the 1991 Green Book is 

suitable for dredged material regulated under 

MPRSA because it is based on biological-effeds 

testing, which takes into account synergistic, 

anbgonistic, and additive effects of all contaminants in the material. This approach includes 

bo+ water-column and benthic impad, and 

assesses both toxicity and bioaccumulation. 

Adaptations of the guidance are also being applied to nearshore and Great Lakes dredge disposal projects, and the tiered testing framework 

may serve as a model for sediment assessmats 

under other regulatory and nonregulatory 

programs. 

13.3.1.1 Suitability for Different Sediment 

OyPS 

Except for extremely coarse- or angulargrain sediments, the tiered-testing approach is 

suitable for all sediment types. The test organisms recommended in the manual are suitable 

for most medium- and fine-grain dredged material. If the dredged material being tested is 

composed of very coarse sediments, or the 

dredged material has other physical properties 

that are potentially incompatible with &ommended test species, alternative organisms may 

be used if they meet 40 CFR 227.27(c) and are 

ecologically relevant to the disposal site. Alternative test organisms may also be necessary 

to avoid grain-shape insensitivities when using 

sediment-ingesting organisms. Noncontaminant-related mortality has been linked on at 

least one occasion to internal organism damage 

that was caused by highly angular sediment of 

moderate grain size (Oakland Harbor sediment; 

Word et al., 1990). Sample handling and 

chemical extraction of very coarse-sediment 

dredged material can also cause analytical 

problems. 

ID general, few analytical problems are 

caused by sediment type. Grain-size problems 

occur rarely because (1) most large-grain-size 

23-13 
Sediment Class+cation Methods Compendium 

sediment contains few contaminants and meets 

the LPC in either Tier I or II, and (2) the 

tiered-testing procedure is relatively flexible 

and allows for alternative evaluation methods. 

13.3.1.2 S&ability for Different Chemicok or 

Classes of Chemical Can&n&ants 

Since the guidance in the 1991 Green Book 

uses effects-based tests, it does not rely on the 

explicit identification of contaminants for decision-making. However, the guidance is suitable 

for detecting and quantifying a wide range of 

organic and inorganic chemicals. In Tier I of the 

testing procedure, target analytes are determined 

for the proposed dredged material. If antamination is suspected, but specific contaminants 

cannot be isolated in the Tier I evaluation, the 

manual recommends that the dredged material be 

scanned for a broad spectrum of contaminants. A 

list of 131 potential target analytes is provided in 

Table 9-l of the 1991 Green Book, Priority Pollutant and 301(l) Pesticides Usted According to 

Structural Compound Class. 

Extensive guidance for laboratory analysis of 

organic and inorganic compounds is provided in 

Section 9 of the manual, Physical Analysis of 

Sediment and Chemical Analysis of Sediment, 

Water, and Tissue Samples. Target analytes for 

the water and tissue analyses are the same as 

those for whole-sediment analyses. Guidance is 

also provided in Section 9 of the manual for 

minimizing salt interferences with the chemical 

analyses. 

13.3.1.3 Suitability for Predicting Eflects on 

Different Organisms 

All four tiers of the tiered-testing procedure 

consider effeds on marine organisms that are 

representative of organisms that are indigenous to 

ODMDSs and have known impad tolerances. In 

Tier I, information on the proposed dredged 

material’s ,effect on laboratory and indigenous 

species is. analyzed. In Tier II, the theoretical 

bioaccumulation potential (TBP) for nonpolar 

inorganic contaminants in the dredged material is 

calculated and compared against that of the reference site. In Tier Ill, water-column toxicity, 

benthic toxicity, and benthic bioaccumulation are 

determined for ecologically relevant laboratory 

organisms. In Tier IV, case-specific bioassays 

and bioaccumulation studies are conduded on 

laboratory and/or field organisms. 

13.3.1.4 Suitability for /n-Place Pollutant 

Control 

The 1991 Green Book was developed to 

determine water-column and bentbic LPC compliance for proposed dredged material, not for 

in-place management of contaminated sediments. 

However, the physical, chemical, and biological 

tests that are recommend&d in the tiered-testing 

procedure are readily adaptable to nondredging 

management of sediments. 

The sediment data that are generated with 

the guidance in the manual must be of sufficiently high quality to develop LPC determinations for the dredged material. If these data 

show that the dredged material does oat meet 

the LPC for ocean disposal, the same data are 

readily adaptable to other sediment-management 

uses, including in-place pollutant management. 

13.3.1.5 Suitability for Source Control 

The purpose of the detailed sampling and 

testing guidance in the 1991 Green Book is to 

fully characterize the dredged material that is 

proposed for ocean disposal. Although it is not 

the intended purpose, this characterization may 

be useful for controlling sources of contaminants 

that are entering the sediments. 

If portions of a proposed project exceed the 

IPC, it benefits the applicant to isolate the compliant and noncompliant areas to economize 

management of the dredged material. For cxample, material that meets the LPC might be disposed of at an ODMDS and material that does 

not meet the IJC might disposed of upland. 

During the process of site characterization, contaminant gradients and source locations might be 

identified (such as occurred in New Bedford 

Harbor, Massachusetts) and remedial or enforcement actions can be directed as appropriate. 

13-14 
I3-“Green Book” Sediment-Testing Approach 

13.3.1.6 Suitability for Disposal 

Applications 

As discussed in Section 13.1 above, the 

guidance in the 1991 Green Book is used to 

conduct LPC evaluations, which are in turn 

used to support ocean-disposal management 

decisions. The manual is not intended to provide guidance on other disposal options available to dredged-material managers. Some 

ocean and nonocean disposal options may require additional or alternative analyses of the 

dredged material to reach decision points. 

Numerous other guidance manuals on dredgedmaterial management are available from 

USEPA and the USACE. 

133.2 General Advantages and Limitations 

13.3.2.1 Ease of Use 

As discussed in Section 13.2.1 above, the 

tiered-testing procedure is relatively flexible. 

The dredged-material evaluator can enter and 

exit the testing procedures at any tier. However, to begin the evaluation in Tier II, III, ‘or IV, 

the data must satisfy the requirements of the 

earlier tier(s). The overall ease of use of the 

testing procedure depends on the evaluator’s 

familiarity with the following: 

n Federal regulations pertinent to 

dredged-material testing and disposal; 

n Sources of existing dredged-material 

(sediment-quality) information; 

m Sampling design; 

n Numerical modeling; 

m Physical, chemical, and biological 

testing; 

l Statistical analysis; and 

m Quality assurance. 

13.3.2.2 Relative Cost 

Tiers I, II, III, and IV are ordered by 

increasing complexity and cost. Tier I is relatively inexpensive and consists solely of assembly and analysis of existing information. 

Tier IV can be very expensive, consisting of 

case-specific toxicity and bioaccumulation 

analysis, including extensive field and laboratory studies. However, significant time and 

resources can be saved if the earlier tiers are 

completed to the maximum extent possibie 

before proceeding to the later tier(s). For example, an in-depth analysis of “grey literature” 

(university reports, etc.) might show the possible existence of ‘hot spots” within a project. 

The sampling plan could then be designed to 

appropriately sample these areas of concern 

during a single sampling event, thereby saving 

the time and expense required to conduct additional sampling at a later time. Similarly, 

money and time will be saved if LPC compliance for nonpolar organic contaminants can be 

shown in the Tier II TBP calculation rather 

than in the Tier III laboratory testing and 

analysis. 

As all dredging ‘projeds contain case-specific components, it is difficult to estimate the 

overall cost of a typical dredged-material analysis. USEPA and the USACE predict that the 

updated methods in the manual would not 

cause a significant increase in evaluation expenses and actually might lead to lower testing 

costs because LPC determinations might be 

achieved earlier in the testing process, thereby 

making full-scale bioassay and bioaccumulation 

laboratory tests unnecessary. Also, as the 

recommended analy!ical methods become refined, market pressures will force costs lower. 

13.3.2.3 Tendency to Be Conservative 

As discussed in Section 13.22 above, the 

tiered-testing procedure is very protective of 

human health and the marine environment. It 

is a sequential and comprehensive analysis of 

the proposed dredged material’s biological 

effects, as shown by previous studies, model13-15 
Sediment Classijicutti Methods Compendium 

ing, and laboratory testing. However, the 

tiered-testing procedure is an “expert system”; 

that is, the product of the procedure (LPC compliance determination) is only as good as the 

information that is integrated into it. 

To reach a defensible and ecologically 

sound LPC evaluation, high-quality information 

is required. There is risk of an inaccurate compliance determination if incomplete or inaccurate information is used, or if good information is misapplied. The regulations and numerous references in the manual should be consulted, and well-trained and experienced evaluators 

should be involved throughout the decisionmaking process. 

X3.3.2.4 Level of Acceptance 

The 1991 Green Book is the official 

USEPA/USACE guidance manual for determining the suitability of dredged material for 

ocean disposal. During the development of the 

updated manual, comments from USEPA and 

USACE Headquarters, USEPA Regions, 

USACE Districts, other federal agencies, port 

authorities, special-interest groups, and the 

general public were solicited, received, and 

addressed as appropriate. In 1990, USEPA and 

the USACE conducted a public meeting on the 

document’ and held six regional training sessions’ on the updated methods. The final 

manual is the product of extensive USEPA/ 

USACE dredged material program experience, 

current state-of-the-art testing methods, and 

review by a wide array of individuals and 

agencies. 

13.3.2.5 Ability to Be Implemented by 

Laboratories with Typical 

Equipment and Handling Facilities 

Many evaluations recommended in the 

1991 Green Book, particularly for organic and 

chemical analysis, require standard laboratory 

‘Washington, DC. 

%amgansett, RI; Gulf Breeze, Q Vicksburg, MS; Newport, OR; San Fran&co, CA; and Washington, DC. 

equipment and handling facilities. However, 

some laboratories have difficulty attaining 

accurate and precise test results for low contaminant concentrations. Agency and contract 

laboratories that presently do not have the 

capabilities to conduct precise analyses will 

have to make significant investments in equipment, personnel, and training. It is expected 

that contract laboratories will choose to specialize in only a few methods to be efficient and 

competitive in the dredged-material testing 

market. Quality assurance (QA) program development, although not equipment-intensive, is 

also a necessary and significant investment for 

testing laboratories. QA programs are necessary to ensure that sample and data integrity 

are of sufficient quality and defensible. 

13.3.2.6 Level of Effort Required to 

Generate Results 

The overall level of effort necessary to 

conduct dredged-material analysis is comparable to that required by the preceding guidance 

(1977 Green Book). The level of effort is 

relatively low in Tier I and relatively high in 

Tiers Ill and IV. 

13.3.2.7 Degree to Which Results Lend 

Themselves to Interpretation 

The analysis of raw data that are generated 

during the tiered-testing procedure is relatively 

complex, especially for bioassay and bioaccumulation test data. Interpretation of results 

is specifically described and decision points 

and values. are clearly defined in the 1991 

Green Book. Section 13 of the manual, Statistical Methods, presents guidance for handling 

the following: 

m Unequal numbers of experimental animals assigned to each treatment container or loss of animals during the 

experiment; 

n Unequal numbers of replications of the 

treatments (i.e., containers or aquaria); 

13-16 
X3- “Green Book ” Sediment-Testing Apprvach 

n Measurements scheduled for selected 

time intervals but actually performed at 

other times; 

n Different conditions of salinity, pH, 

dissolved oxygen, temperature, etc., 

among exposure chambers; and 

m Differences in placement conditions of 

the testing containers or in the animals 

assigned to different treatments. 

USEPA and the USACE are presently developing software and additional guidance to facilitate data interpretation for dredged-material 

evaluations. 

13.3.2.8 Degree of Environmental 

Applicability 

The USEPA/USACE (1991) effects-based 

approach used to evaluate marine sediments has 

wide environmental and regulatory applicability. The approach uses test organisms that 

n Are sensitive to impact; 

l Are reasonable representatives of indigenous ODMDS species; 

n Fulfill the species categories required 

by 40 CFR 227.27(c,d); 

n Have extensive test databases; and 

n Are hardy enough to withstand laboratory procedures. 

Alternative test species that meet the guidance 

in the. 1991 Green Book may be used to avoid 

testing problems such as grain-size tolerance 

and seasonal availability. Complete elucidation 

and quantification of all chemical components 

in the sediment are useful, but not required, for 

regulatory decision-making. The overall 

approach is environmentally conservative and 

relatively economical. 

One feature of the 1991 Green Book guidance posing environmental limitations is the 

numerical modeling that is used in Tier I and II 

water-column evaluations. The ADDAMS 

models are not suitable for calculating watercolumn impacts at disposal sites that are 

extremely shallow (i.e., where the discharge 

period from the disposal vessel is longer than 

the descent time to the bottom). Additionally, 

there is some uncertainty about the applicability of the models for extremely deep (a200 m) 

ODMDSs. 

13.3.2.9 Degree of Accuracy and Precision 

“Ihe 1991 Green Book guidance strongly 

emphasizes the importance of a comprehensive 

QA program to. achieve sufficient data quality 

during the tiered evaluation process. QA issues 

are addressed in subsections throughout the 

data-generation sections of the manual, and 

Section 13, Quality-Assurance Consideration, 

gives guidance on the structure and components of QA programs and data-quality 

assessment. 

The general guidance for QA program development includes information on field and 

laboratory sample handling, personnel training, 

and documentation. For chemical analyses, the 

guidance recommends appropriate use of ~nethod blanks, procedural blanks, matrix 

spike/matrix-spike duplicates (MSSD), and 

standard reference materials (SRM) to detamine accuracy and precision of the data. For 

biological testing, the importance of controlsediment tests, reference-site tests, and reference-toxicant testing is discussed. 

WA fiTAlUS 

WA.1 Extait of Use 

Tbe1991GreenBookguidancewillbeap 

plied to all evaluations for dredged material &at is 

pfopo6e4f for disposal outside the baseline of the 

tenitorlal se8 (non-state waters). Until completion 

of ongoing work on a national testing manual fa 

disposal shoreward of the baseline of the tenitorial 

13-17 
Sediment Classificution Methods Compendium 

sea (Ckan Water Act section 404 waters), portions 

of the Green Book guidance are also expected to 

be applied to nearshore and internal-water 

dredged-material disposal projects in the United 

states. 

13.4.2 Extent to Which the Approrcb Has 

Been Field-Valida@d 

Large potions of the tiered-testing procedure 

for dredged material have been field-validated 

since the publication of tbe original guidance in 

1977 by ongoing state and federal dredging programs. Several large-scale, long-term 

USEPAAJSACE projects in the New England and 

West Coast regions have applied and improved on 

the methods in the 1977 manual. ‘2be guidance in 

the 1991 Green Book contains methods proven for 

marine sediment analyses, developed for national 

testing consistency, and organized into tiers for 

efficient compliance determination. ‘Ibe tiered 

approach for environmental monitoring of aquatic 

ecosystems is strongly recommended by the 

National Research Council (NRC, 1990). 

13.43 Reasons for IAmited Use 

Only extreme time and resource constraints 

(national emergencies, etc.) would limit the use of 

the guidance in the manual. Most of the recommended procedures are already widely applied. 

13.4.4 Outlook for Future Use sad 

Development 

USEPA and the USACE will continue to 

support and apply the guidance in the manual both 

nationally and regionally. Ongoing public and 

private research and development of evaluation 

methods will continue to expand federal and state 

dredging-program experience. 

?be manual will be revised at a future date 

based on (1) the findings of an EPA SAB review 

(SAB, 1992); (2) technical advances in assessing 

sediment contamination and marine environmental 

impact; and (3) changes to the Qcean Dumping 

Regulations. 

135 REFERENCES 

NRC. 1990. Managing troubled waters: The 

role of marine environmental monitoring. 

National Research Council. National Academy Press, Washington, DC. 125 pp. 

MB. 1992. An SAB report Rwiew of a 

testing manual for evaluation . of dredged 

material proposed for ocean disposal. Prepared by the Sediment Criteria Subcommittee of the Ecological Processes and Effects 

Committee; USEPA Science Advisory 

Board, Washington, DC. EPA-SAB-EPEC92-014. 

USEPA/USACE. 1977. Environmental Protection Agency/United States Army Corps of 

Engineers Technical Committee on Criteria 

for Dredged and Filled Material. Ecological 

evaluation of proposed discharge of dredged 

material into ocean waters; Implementation 

manual for section 103 of Public Law 92- 

532 (Marine Protection, Research, and Sanctuaries Act of 1972). July 1977 (second 

printing April 1978). Environmental Effects 

Laboratory, United State& Army Engineer 

Waterways Experiment Station, Vicksburg, 

MS. 24 pp + appendices. 

USEPA/USACE. 1991. Environmental Protection Agency/United States Army Corps of 

Engineers. Ecological evaluation of proposed discharge of dredged material into 

ocean waters. January 1990. United States 

Environmental Protection Agency, Office of 

Marine and Estuarine Protection, Washington, DC 20460. USEPA-503-8~90/002, 

219 pp + appendices. 

Word, J.Q., JA. Ward, JA. Strand, N.P. Kahn, 

and A.L Squires. 1990. Ecological evaluation of proposed discharge of dredged 

material from Oakland Harbor into ocean 

waters (Phase II of 42Foot Project). Repared for United States &my Corps of 

Engineers. U.S. Department of Energy 

Contract No. DE-AC%-76RLO 1830. September 1990. 

13-18 
CHAPTER 14 

National Status and Trends Program 

Approach 

Edward R. Long 

Coastal Monitoring and Bioeffects Assessment Division 

National Oceanic and Atmospheric Administration 

7600 Sand Pt. Way, NE, Seattle, WA 98115 

(206) 526-6338 

Donald D. MacDonald 

MacDonald Environmental Sciences, Ltd. 

2376 Yellow Point Road, R. R. #3, Ladysmith, SC, Canada VOR 2E0 

Sediment quality criteria based on multiple 

methods have been recommended for broad 

applications in the United States (USEPA/SAB, 

1989; Adams et al., in press). The approach used 

by the National Status and Trends Program 

(NSTP) of the National Oceanic and Atmospheric 

Administration (NOAA) to develop informal, 

effects-based guidelines involves the identification 

of the ranges in chemical concentrations associated with biological effects based on a weight of 

evidence from many studies. In this approach, the 

data for many chemicals are assembled from 

modeling, laboratory, and field studies to determine the ranges in chemical concentrations that 

are rarely, sometimes, and usually associated with 

toxicity. The data from many of the studies of the 

individual approaches described elsewhere in this 

document are compiled and examined to develop 

no-effects, possible-effects, and probable-effects 

ranges (Figure 14-1). 

14.1 SPECIFIC APPLICATIONS 

14.1.1 Current Use 

The NSTP Approach was used initially to 

develop informal guidelines for use by the National Status and Trends (NS&T) Program (Long and 

Morgan, 1990; Long, 1992). NOM analyzes 

sediments from numerous locations nationwide as 

a part of its monitoring program. The guidelines 

were developed as tools for identifying locations 

in which there is a potential for toxicity to living 

resources for which NOAA is the federal steward. 

Areas in which chemical concentrations often 

exceeded the guidelines were identified as high 

priorities for investigations of toxicity with biological tests. 

Environment Canada evaluated many candidate approaches to the development of sediment 

quality guidelines and elected to develop its 

national guidelines using the NSTP Approach 

(MacDonald and Smith, 1991; MacDonald et al., 

1991). The Florida Department of Environmental 

Regulation elected to use the NSTP Approach to 

develop state sediment quality guidelines as a part 

of its sediment management strategy (MacDonald, 

1992). The California Water Resources Control 

Board will use the NOM guidelines in its initial 

evaluations of ambient chemical data. Following 

that step, data from field studies, laboratory 

bioassays, and equilibrium partitioning models 

will be used to develop sediment -quality objectives (Lorenzato et al., 1991). Finally, the International Council for Exploration of the Sea Study 

Group on the Biological Significance of Contaminants in Marine Sediments has elected to adopt 

the NSTP Approach in the development of guidelines for participating nations (Dr. Herb Windom, 

Working Group on Marine Sediments, ICES, 

personal communication), 

Guidelines developed with the NSTP Approach were used by NOAA to identify chemicals 

that occurred in concentrations that were sufficiently high to warrant concern and to identify 

sampling sites and areas in which there was a 

potential for toxicity (Long and Morgan, 1990; 
Sediment Classification Methods Compendium 

100 

Frequency Distribution 

of Effects (percentile) 

50 

10 

0 

Possible Effects Probable Effects 

ERL ERM 

Chemical Concentration 

Figure 14-1 Conceptual outline of the relationship between the NSTP Approach guidelines and the no-effects, possible effects, and 

probable-effects ranges in chemical concentrations. 

14-2 
1ANSTP Approach 

Long et al., 1991; Iong and Markel,. 1992). It 

was presumed that the potential for toxicity was 

relatively high in areas where numerous chemicals 

exceeded the upper bounds of the guidelines. 

Likewise, it was assumedthat the potential for 

toxicity was relalively low in areas where none of 

the chemical concentrations exceeded the lower 

bounds of the guidelines. In those regions with 

the highest potential for toxicity, NOAA has 

implemented regional surveys of toxicity; using a 

battery of biological analyses and tests. 

Also, NOAA has used the guidelines in 

assessments and prioritization of hazardous waste 

sites (Dr. Alyce Fritz, NOAA Hazardous Materials 

Response and Assessment Division, personal 

communication). Other agencies and consultants 

have used the guidelines as a means of placing 

ambient chemical data into perspective with 

respect to the potential for toxicity (for example, 

Squibb et al., 1991 for New York/New Jersey 

Harbor; Mannheim and Hathaway, 1991 for 

Boston Harbor; Soule et al., 1991 for Marina Del 

Rey). The Florida Department of Environmental 

Regulation has used the guidelines as informal 

tools for interpreting ambient chemical data and 

for identifying regional priorities for sediment 

quality management (MacDonald, 1992). 

14.13 Potential Use 

Potential uses of the guidelines are as follows: 

B Identification of potentially toxic chemicals in ambient sediments; 

n Ranking and prioritization of areas and 

sampling sites for further investigation; 

n Assessment of potential ecological hazards of contaminated sediments; 

n Design of spiked-sediment bioassay experiments; 

n Description of the kinds of toxic effects 

previously associated with specific concentrations of chemicals; 

n Quantification of the relative likelihood of 

toxicity mer specific ranges in chemical 

concentrations; and 

W Identification of the need for sediment 

management initiatives. 

14.2 DESCRXPTION 

14.2.1 Description of Method 

The N!STP Approach involves a simple evaluation of available data to identify three ranges in 

concentrations for each chemical: 

l No-Effects Range: The range in concentrations over which toxic effects are rarely 

or never observed; 

B Possible-Effects Range: T&e range in 

concentrations over which toxic effe& 

are occasionally observed, and 

n Probable-Effects Range: The range in 

concentrations over which toxic effects 

are frequently or always observed. 

These ranges are identified by evaluating 

information from numerous studies in which 

matching biological and chemical data were 

developed. The specific steps in the method are: 

(1) Compile matching chemical and biological data from laboratory spiked-sediment 

bioassays, equilibrium-partitioning models, and field studies and determine the 

chemical concentrations associated with 

no observed effects and those associated 

with adverse effects. 

(2) Enter the data into a database, including 

the type of biological test performed, the 

adverse effect(s) measured, the chemical 

concentrations associated with observations of either effects or no effects, the 

type of study method and approach, and 

the degree of concordance between the 

14-3 
Scdimen t Classification Methods Compendium 

measure of effects and the concentration 

of the chemical. 

(3) For those analytes for which sufficient 

data exist, prepare data tables sorted 

according to ascending chemical concentrations. 

(4) Arithmetically determine the no-effects 

range, possible-effects range, and probable-effects range for each chemical. 

The steps taken to select and screen candidate data sets are described in Section 

14.2.1.2.3. The approach is intended to encourage periodic updates as new data become available. 

Two slightly different methods have been 

used to determine the three chemical ranges. 

First, two percentiles in the chemical concentrations associated with toxicity were derived by 

Long and Morgan (1990): the lower 10th percentile and the 50th percentile (median). The 

lower 10th percentile was identified as the 

Effects Range-Low (ERL), and the median was 

identified as the Effects Range-Median (ERM). 

In their evaluation of the ascending data tables, 

Long and Morgan (1990) used only the chemical 

concentrations that had been associated with 

toxicity (i.e., the “effects” data). The conceptual 

basis for this approach and the three ranges are 

illustrated in Figure 14-2. 

Later, MacDonald (1992) identified the three 

ranges with a method that used both the concentrations associated with biological effects (the 

“effects” data) and those associated with no 

observed effects (the “no-effects” data). In this 

method, a threshold effects level (TEL) was 

calculated first as the square root of the product 

of the lower 15th-percentile concentration associated with observations of biological effects 

(the ERL) and the 50thpercentile concentration 

of the no-observed-effects data (the NER-M). A 

safety factor of 0.5 was applied to the TEL to 

define a No-Observable-Effects Level (NOEL). 

Next, a Probable-Effects Level (PEL) was 

calculated as the square root of the product of 

the SOth-percentile concentration of the effects 

data (the ERM) and the 85th-percentile concentration of the no effects data (the NER-M). 

Neither of these methods is preferred or 

advocated over the other. The significant feature of this approach is the use of a weight of 

evidence developed in the ascending tables, not 

in the specific method of using the da& tables. 

In addition to the two metbods described here, 

many others could be applied to the ascending 

data tables to derive guidelines. The method 

used by MacDonald (1992) considered both the 

“effects” and “no-effects” data, whereas that of 

Long and Morgan (1990) used only the “effects” 

data. Different percentiles in tbe ascending data 

were used in the two methods. Despite these 

differences in the methods, the agreement between the NOELs and ERLs and between tbe 

PELS and the ERMs was very good, usually 

within a factor of 2. 

In both documents, the lower of the two 

guidelines for each chemical was assumed to 

represent the concentration below which toxic 

effects rarely occurred. The range in concentrations between the two values was that in which 

effects occasionally occurred. Toxic effects 

usually or frequently occurred at concentrations 

above the upper guideline value. 

As an example, ,Figure 14-2 compares the 

frequency distribution of toxic effects and noeffects data associated with concentrations of 

napthtbalene to the ERL and ERM concentrations for napbthalene. Long and Morgan (1990) 

reported the ERL as 340 ppb dry wt. and the 

ERM as 2100 ppb dry wt. for naphthalene, 

based on an ascending data table of 49 data 

points. These guidelines defined three ranges of 

chemical concentrations: the no-effects range 

(O-340 ppb); the possible-effects range 

(340-2100 ppb); and the probable-effects range 

(~2100 ppb). Only 10.5 percent of the chemicat 

concentrations below the ERL were associated 

with toxic effects; suggesting that toxicity is 

unlikely below the ERL concentrations. In 

contrast, 81 percent of the chemical concentrations between the ERL and ERM values were 

associated with the toxic effects and 93 percent 

of the data points were associated with toxicity 

at concentrations above the ‘ERM value. 
No Bffects Possible Effects 

0 340 2100 

Naphthalene (ppb) 

flgun 14-2 hquency dlstributhw of naphthalene concentration wsodated with toxic effects below the ERL value, between the ERL 

end ERM v&es, nd above the ERM vdua (from hg l nd Morgan, looo). 
Sediment Classification Methods Compendium 

14.2.1.1 Objecfives and Assumptions 

The objective of the NSTP Approach is to 

provide informal, effects-based guidelines that 

are based on a weight of evidence and reported 

as ranges in concentrations. The guidelines are 

based on chemical concentrations associated 

with measures of biological effects, thereby 

providing toxicological and/or biological relevance to the guidelines. They are based on data 

from multiple studies and research methods, thus 

providing a weight of evidence. In recognition 

of the variability in the kinds of data that are 

available, they are presented as ranges, instead 

of absolute values, thereby providing a flexible 

interpretive tool with broad applicability. They 

are presented along with all of the supporting 

evidence in ascending tables, providing tbe user 

an interpretive framework for comparison with 

ambient data. 

In this approach it is assumed that tbe data 

from all individual studies are equal in weight and 

credibility, although they may have involved very 

different methods and test endpoints. It is assumed that the methods used by the individual 

investigators were reasonably accurate. Most 

important, it is assumed that as the concentrations 

increase, the potential for toxicity also increases, 

thereby providing a conceptual basis for identifying the ranges in concentrations frequently associated with no toxic effects and those frequently 

associated with toxic effects. The guidelines can 

be formulated to account for site-specific factors 

that control bioavailability (see Section 14.3.1.1). 

14.2.1.2 Level of Eflort 

14.2.1.2.1 Type of Sampling Required 

The NSTP Approach relies on the use of a 

database compiled from a wide variety of sediment quality assessments. The database currently 

contains over 800 entries generated by the three 

major approaches to the establishment of effectsbased guidelines: equilibrium-partitioning models; 

laboratory spiked-sediment bioassays; and various 

assessments of matching, field-collected, sediment 

chemistry, and biological effects data, The NSTP 

Approach was specifically designed to use existing 

data, therefore eliminating or minimizing the need 

for additional sampling. However, evaluation of 

the regional applicability of the guidelines could, 

in some cases, require further site-specific investigations, the magnitude of which could vary considerably. 

14.2.1.2.2 Methods 

The methods for deriving numerical sediment 

quality guidelines using the NSTP Approach are 

summarized in Section 14.2.1. Also, these methods are described by Long and Morgan (1990) and 

MacDonald (1992). 

14.2.1.2.3 Types of Data Required 

The NSTP Approach was intended to integrate 

a diverse assortment of information into a single 

database to support the derivation of numerical 

guidelines. Consequently, data from numerous 

modeling, laboratory, and field studies were 

collated into one database. Ideally, the database 

used to establish guidelines should include entries 

from all three of these types of approaches. 

Suitable data were available from a wide variety 

of sources. While collection and analysis of these 

data sets were labor-intensive, subsequent, inuemental updates of the database should be relatively simple and inexpensive. 

The data compiled from numerous studies 

were entered into the Biological Effects Database 

for Sediments (BEDS) by MacDonald (1992). All 

of the compiled data were fully evaluated prior to 

incorporation into the BEDS to ensure internal 

consistency in the database. The screening procedures used to support the development of the 

3EDS were designed to ensure that only relevant 

and highquality data were used to derive the 

guidelines. No subjective biases were employed 

in screening the data; as many sources of data 

were included as possible. Candidate data from 

each study were evaluated to determine the acceptability of the experimental design, the test 

protocols, the analytical methods, and the statistical procedures that were used. Only data in which 

there were matched measures of sediment chemjs- 
IANSTP Approach 

try and biological effects were included. The 

database included only those data in which either 

statistically significant biological results were 

obtained or in which major differences in the 

biological results between samples were reported. 

The BEDS currently includes over 800 data 

entries, mainly data from studies performed 

throughout North America. It was developed 

jointly by NOAA, Florida Department of Environmental Regulation, Environment Canada, and 

MacDonald Environmental Services L&l. 

In the evaluation of candidate data from field 

studies, only those data were used in which at 

least a lo-fold difference in the concentrations of 

at least one chemical among the samples was 

reported. Once this criterion was met, the data 

from many of the field studies were evaluated to 

determine the mean chemical concentrations in 

toxic samples (i.e., significantly different from 

controls) and those in nontoxic samples or in 

samples with relatively depauperate benthic cornmunities (i.e., those with low abundance or species 

richness) versus those with more robust communities. Further, those mean concentrations in biologically affected samples that exceeded by twofold 

or more the mean concentrations in the background, reference, or nonaffected samples were 

assigned an asterisk in the ascending tables. The 

asterisks symbolized that a biological effect was 

noted and that there was a strong association 

between the chemical gradient and the biological 

gradient. Concentrations associated with nontoxic 

reference conditions were noted as “no effects.” 

Those in which there was no concordance between 

the measures of effects and chemical concentrations were noted as “no gradient” or “no concordance.” The concentrations derived in the modeling and spiked-sediment bioassays were always 

assigned asterisks. The concentrations with 

asterisks were used as ‘effects” data by both Long 

and Morgan (1990) and MacDonald (1992). 

14.2.1.2.4 Necessary Hardware and Skills 

l’be primary skills required to derive guidelines 

are associated with the development of the database. 

Expertise is required to evaluate the suitability of the 

biological and chemistry data, using the screening 

criteria. This process requires experience in the 

evaluation of sediment data and the methods that 

were used to develop the data. 

The database has been developed on a personal 

computer and is readily transferable to other systems, but requires knowledge of the use of a oomputer. ‘l’he database provides a means of storing 

and accessing all of the information that relates 

chemical concentrations to adverse biological effeds. This information can be manipulated in Uris 

environment or exported into other formats. 

14.2.1.3 Adequacy OfLlocrmrenti 

The NSTP Approach was documented by Lrng 

and Morgan (1990), in which the approach ‘was 

peer-reviewed both within and outside NOA& A 

second printing of the document was issued in 1992, 

following further review. A synopsis of the ap 

preach was descrii in a scientific journal (Long, 

1992). The approach has been descrii orally in 

numerous technical and scientific forums. MacDonald and Smith (1991) and MacDonald et cll. 

(1991) desaii the application of the approach in 

the development of guidelines for Canada. MacDonald (1992) descrii the use of the approach in 

a statewide sediment management strategy for 

FlOriQ. 

14S.2 ApplkabWy of Method to H-II 

Health, Aquatic I.&z, or WMlife 

Px-dedhn 

‘Ibe guidelines are intended to povide an estimate of the potential for adverse biological effects of 

sediment-associated contaminants ar benthic organ- 

~ba&onaweightofevi&ncefromanalyses 

performed with multiple species and/or biological 

communitie5. They accomm odate and rely on the 

data from te& of acute and chronic toxicity and liwn 

analyses of bent& community structure. The guidelinesareba&ondatafmmmanydifferezttareasand 

oceanographic regimes, thereby broadeking their 

applicability. Currently, the data entered into the 

BEDS are from atly marine and estuarine areas. 

pe guidelines prwide a means of numticai~y 

estimating the percent frequency of biological effects 

over the lhree ranges of concentrations. The axxnd14-7 
Sediment Ciassijication Methods Compendium 

ing tables accunpanying the guidelines also provide 

a supplementary basis for interptiing new ambient 

cbtical data. Ako, these tables provide a visual and 

statistical means of estimating tbe relative degree of 

certainty in lhe guidelines. 

‘I%e guidelines are not intended to be used for the 

prokction of human life or wildlife. Rather, they are 

intended to be used in est*ating the potential for 

adverse effects among benthic communities. 

14.23 AbllityofMetbodtoGenmtc 

Numdcd Crittxia for Specifk 

Chemkak 

kg and Morgan (1990) reported numerical 

guidelines for 41 chemicals, including 12 trace 

metals, 18 polynuclear aromatic hydrocarbons 

(PAHQ, and 11 synthetic organic compounds. 

MacDonald (1992) developed guidelines for 9 trace 

metals, total PCS, 13 PAHs, 3 classes of PAHs, and 

2 pesucides. 

Grmplually, guidelines derived using this 

approach could be developed for any toxic chemical, 

provided sufficient data exist and provided the 

toxicity of the ckrnical is doseresponsive. Long and 

Morgan (1990) assigned a high degree of a&idence 

to gui&lines for ckmicals for which data existed 

Can many different approaches, different regiq 

and in which there was a good agreement in the data 

hn different studies. MacDonald (1!392) calculated 

guidelines only for those chemicals for which there 

was a minimum of 40 data points, after determining 

the minimum amount of data necessq to calculate 

reliable and consistent values. These minimum data 

requirements were established by iteratively calculating guidelines using data sets of increasing size (e.g., 

4 to 60 data points) and determining when the 

estimate of the guidelines stabilized. 

143 USEFUTAE!S 

143.1 Ewiro~~ntal Applicability 

143.1.1 Sui&biLity for Diferent &xGnent lopes 

The NSI’P Approach can be applied equally to 

any sediment type that occurs in fkshwater, estuarine, 

and marine environments. Since the database that 

supports the guidelines amtains information fian a 

wide variety of sediment types, the resultant guide 

lines are considered to be widely applicable. An 

increasing amount of information suggests that the 

bioavailabitity, and, therefore, toxicity, of many 

contaminantsisconlrolledbysu&factorsasTOC, 

AVS, and grain size. ‘Ibe BEDS aa~ently accunmodatestbedarafortbesevariable&;and,auW 

quemtly, the guidelines could be- namalized to the 

appropriate faders that a~trol bioavailability. 

Hawever, insufficjent information cum~Iy exists to 

derive guidelinea~that are expressed in these terms. 

It is anticipated that future revisions of the guidelines 

will be expressed in lhese temq thereby increasing 

their applicability. 

Partly to increase the &ability of the guidelines 

to different sediment types, they are eqessed as 

ranges in owcentralions, not absolutes. These ranges 

povide a basis for evaluating chemical concentratks 

in the different types of sediments repreamted in the 

BEDS. In addition, the asamding data tables used to 

generate the guidelines can be examined to calculate 

frequency distriiutions of effects and no effecis 

within each range of concentrations. Ikse fkquency 

dislriiutions an be used as estimates of the pfobability of toxic effeds. 

14.3.X.2 Suhbility for Differeent Chemicals 

or Classes of CkmicaLr 

The approach can be applied to a wide variety 

of chemicals for which analytical methods are 

available. Thus far, numerical guidelines have 

been developed by hng and Morgan (1990) and 

by MacDonald (1992) for 43 and 28 chemicals or 

classes of chemicals, respectively. Data are 

included in the BEDS for over 200 chemicals or 

classes of chemicals. Guidelines could be developed for all of these substances when sufficient 

information becomes available. 

14.3.1.3 Suitability for Predicting Eficts on 

Diflerent Organisms 

Since the database compiled from many 

different studies is based on tests or analyses 

performed with many different species, the guidelines are widely applicable ti benthic organisms. 
1ASTP Appronch 

In addition, the species studied in each investigation is(are) listed in the database; therefore, 

species-specific applicability can be evaluated by 

the users. Furthermore, the ERL values often are 

based on data from relatively sensitive species or 

life stages, and, therefore, can be used as guidelines suitable for the protection of sensitive species. 

14.3.1.4 Suitability for In-Place Polkant 

Control 

Numerical sediment guidelines developed 

using the NSTP Approach can be used in a variety 

of ways as a tool in pollutant control. Specifically, these assessment tools respond to regulatory 

requirements by: 

n Providing a basis for evaluating existing 

sediment chemistry data and ranking areas 

of concern and chemicals of concern in 

terms of their potential for causing toxicity and 

n Identifying the need for further investigations, such as biological testing, to support regulatory decisions. 

As is the case with all of the other approaches 

that rely on data collected in the field, the guidelines derived using the N!XP Approach integrate 

information obtained from studies of complex 

mixtures of contaminants and thereby consider 

their interactive effects. Consideration of the 

effects of contaminant mixtures is an advantage in 

the assessment of in-place pollutants in real-world 

conditions. However, this approach also relies on 

and gives equal weight to the data from equilibrium-partitioning models and laboratory spikedsediment bioassays performed with single chemicals (see Section 14.2.1.1). 

14.3.1.5 Suitability for Source Control 

A reasonable amount of confidence in sediment quality guidelines is needed to justify using 

them in source control adions. Since the guidelines are developed with a weight of evidence 

compiled from many different studies, they provide a credible and defensible basis for evaluating 

contaminants in real-world conditions. Ihe 

guidelines provide an efficient basis for identifying priority chemicals and priority areas that 

would benefit from source controls. In addition, 

the ascending tables provide a basis for estimating 

the probability of observing adverse effects at sites 

of interest, reducing the probability of effeds 

through source controls, and evaluating the improvements in sediment quality following the 

implementation of source control measures. 

14.3.1.6 Suitability for Dredged Material 

Disposal Applications 

Neither the numerical guidelines nor the 

framework that have been developed for their 

application are intended to replace aaxpted testing 

protocols for dredged material disposal evaluations. Nonetheless, these guidelines can provide 

relevant tools for estimating the potential for 

adverse biological effects of contaminants associated with solid-phase sediments. 

143.2 Gemen Advmhges l d Dludvmhgee 

14.3.2.1 Ease of Use 

The approach has the advantage of relying on 

existing data. Tberefore, guidelines am be devtloped relatively quickly and easily. 

‘The original efforts by Long and Morgan (1990) 

and MacDonald (1992) to assemble the databasea 

used to develop the guidelines were labas-intemive. 

Numerous reports and data sets were located, md a 

huge amount of data was entered into spreadshe 

However, these data now exist in a central&d, 

computerized database, the BEDS. Subsequent 

dexivations of guidelines based on iterative upansions of the BEDS database should be relatively 

quick, easy, and inexpensive. 

‘I%e guidelines are easily used and interpr&d 

@xxkal data can be readily cunpnred with the 

guidelines and with the -ding tables. The fre 

quemy of occurrena of toxicity over the no-&f* 

po&ble+ffects, and probable-effects ranges can be 

calculated and compared with the chemical data. 
Safiment Classification Methods Compendium 

Sediments in which numerous chemicals occur at 

concentrations that fall within the probable-effects 

ranges have a higher probability of being toxic than 

those in which most of the chemical concentrations 

are within the no-effects range. This type of simple 

interpretation makes the guidelines very easy to use. 

14.322 Relatiw Cost 

The original effart of Long and Morgan (1990) 

involved roughly one year of labor. “fbe confinnation and expansion of the database by MacDonaId 

(1992) involved more than another year of labor. 

The costs of subsequent iterations of the guidelines 

based on further expansions of the database would 

vary with the amount of data entered and the number of chemicals. The calculations of the guideline 

values themselves are very simple and quick. Also, 

the guidelines can be used very quickly and easily. 

If the necessary data are not available for entry 

into a database, then the costs to generate them 

could be relatively high. If initiated de nouo, 

modeling, bioassay, and field studies necessary to 

generate sufficient data could vary considerabIy in 

costs and time, depending on the amount of data 

needed. 

14.23 Tendency to Be Cmservative 

The predictive capabilities of the guidelines 

have not been independently quantified. The 

protectiveness of the guidelines could be increased 

by considering data only from chronic sublethal 

endpoints or by applying a numerical safety factor, 

such as was applied in the Florida guidelines (MacDonald, 1992). Also, the guidelines would become 

more conservative if data were included only from 

areas in which toxicants were highly bioavailable. 

14.3.2.4 Level of Acceptance 

7he NSTP Approach has been published by 

NOAA, following an in-house and outside peer 

review. It has been published in a peer-reviewed 

scientific journal. The approach has been used by 

Environment Qnada and Florida Department of 

Environmental Regulation in the development of 

their respective guidelines. It has been adopted by 

a committee of the International Council for Exploration of the Sea for use by member nations. ‘Xbe 

State of California has adopted a similar approach to 

the development of sediment quality objectives 

(Imenzato et al., 1991). 

The numerical guidelines developed by use of 

the approach have been used by NOAA to compare 

and rank the potential for toxicity at monitoring sites 

nationwide, within San Francisco Bay, and within 

Tampa Bay. Approximately &O copies of the 

report by Long and Morgan (1990) have been 

distributed. Users of the report have compared 

ambient concentrations with the guidelines in 

assessments of hazardous waste sites, analyses of 

prospective dredge material, evaluations of survey 

and monitoring data., and estimates of ecological risk 

(for example, Mannheim and Hathaway, 1991; 

Soule et uf., 1991; Squibb et al., 1991). NOM 

routinely uses the guidelines in its estimates of 

ecological risk at National priority List hazardous 

waste sites. ‘lhe guidelines have been used as a 

basis for interpretation of chemical data in court 

cases. 

14.3.2.5 Ability to Be Implemented by 

Laboratories with Typical 

Equipment and Handling Facilities 

The spreadsheets and database needed to 

generate the guidelines can be prepared with a 

personal computer and need not be very complicated. Entry of data into the database and the 

generation of the asending tables are very simple. 

The calculations of the guidelines can be performed manually, on a desk-top calculator or a 

personal computer. IIre database can be supplemented with new data as they become available. 

Implementation of the approach can become more 

laborious and complicated if the necessary data 

must be generated de rwvo. 

24.3.2.6 Level of Effort Required to 

Generate Results 

As outlined in Section 14.3.2.2, the level of 

effort required in the development of the original 

set of guidelines was relatively bigh. Subsequent 

iterations of the guidelines for other purposes, 

14-10 
lA?JSTP Apoach 

other chemicals, or for the same chemicals following additions to the database would be relatively 

f=Y* Entry of new data points from spikedsediment bioassays, equilibrium-partitioning 

models, or apparenl effects thresholds into the 

database would require only a few minutes. 

Manipulation of raw matching data from biological and chemical analyses performed in a field 

study would require from a few hours lo several 

days, depending on the size of the data set, followed by entry of the data points into the database. 

14.3.2.7 Degree to Which Resulfs Lend 

Themselves to Interpretation 

The guidelines and the ascending data tables 

on which they are based can be used in a number 

of ways. First, the data from analyses of ambient 

samples can be compared visually with the two 

numerical guidelines to determine whether the 

ambient concentrations exceed either of the guidelines. Second, the ambient concentrations can be 

compared with the data in the ascending tables to 

determine the kinds of toxic effects that have been 

observed in previous studies at tbe concentrations 

of concern. Finally, the frequencies of toxicity in 

the no-effects, possible-effects, and probableeffects ranges can be used to predict the probability of toxicity associated with any contaminant 

concentration. 

The guidelines developed thus far with this 

approach do not account for the effects of factors 

that control bioavailability of the toxicants. This 

is not a weakness of the approach; rather, it is a 

weakness of the available data. Nevertheless, this 

wealmess may hinder interpretation of ambient 

data with the guidelines, The BEDS database 

includes a provision for entering data from analyses of acid volatile sulfides and total organic 

carbon (and other potential normalizers) and, 

therefore, would lend itself to recalculation of 

guidelines normalized to these factors once the 

necessary data become available. 

An important strength of this approach is that 

it provides the user some flexibility in the use and 

interpretation of the guidelines. All of the data 

are provided in ascending order for the user to see 

and evaluate. The degree of certainty in the data 

can be assessed and judged by the user. Ranges 

in concentrations are provided, instead of rigid, 

single absolute values. 

One of the most attractive features of this 

approach is the estimation of the probability of 

biological effects, based on the frequency distributions of effects for each chemical. For example, 

the data in the BEDS database indicate that only 

5.8 percent of the chemical concentrations within 

the no-effects range for cadmium (0 to 1 mg/kg) 

determined by MacDonald (1992) were associated 

with adverse biological effects (Figure 14-3). 

These data suggest that there is a low probability 

of observing adverse effects within this range. 

Within the probable effects range for cadmium 

(~7.5 mg/kg), roughly 68 percent of the database 

entries were associated with adverse effects. 

These data suggest that there is a relatively high 

probability of observing adverse effects within this 

range. Positive concordance between frequency of 

effects and chemical concentrations should inspire 

confidence in the guideline values. 

Evaluation of the guidelines for mercury 

reveals that a lower level of confidence should be 

placed on the guidelines for this element. The 

data in the BEDS database indicate that within the 

no-effects range (0 to 0.1 mg/kg), roughly 7 

percent of the entries were associated with adverse 

effects (Figure 14-4). However, frequency distributions of effects are similar within the possibleeffects range (0.1 to 1.4 mg/kg) and the probableeffects range (~1.4 mg/kg), namely 30.1 percent 

and 33.3 percent, respectively. Therefore, it is 

more difficult to adequately determine the unacceptable levels of mercury in sediments than with, 

say, cadmium. 

14.3.2.8 Degree of Environmental 

Applicability 

The guidelines are highly applicable to the 

interpretation of environmental data. They are 

generated with data from environmentally realistic 

field studies, as well as theoretical modeling 

studies and controlled laboratory experiments. 

They are generated with data from many differeat 

regions in which the mixtures and concentrations 
Sediment Classifktion Methods Compendium 

% of data > or < concentration 

60 - 

No Effects Possible 

Effects 

50 - 

5.8% Hits 

’ 26% Hits 

40 - 

30 - 

20 - 

10 - 

0 ! ’ “““’ I I I '1"' I I 'III 

0.01 0.1 1 

Concentration 

- No Effects -t - Effects 

Probable Effects 

68.2% Hits 

+ 

I” I I I ""'1 I I "'I( 

10 1.00 1 

(ug/d 

00 

14-12 
IAhrSTP Approach 

60 

50 

40 

30 

20 

10 

0 

% of data > or < concentration 

No Effects 

7% Hits 

Possible Effects 

30.1% Hit+” 

. 

Probable Effects 

33.3% Hits 

0.1 1 10 100 

Concentration (ug/g) 

- No Effects -‘Effects 

M-z3 
Sediment Classification Methods Compendium 

of chemicals differ and in which sedimentological 

properties differ. They are generated with tests 

using differeni species with different sensitivities 

to toxicants. They are universally applicable in 

Nortb America since they were generated with 

data from many regions in the United Stales and 

Canada. Confidence in the utility of the guidelines is inspired by the weight of evidence from 

these multiple studies. 

14.3.2.9 Degree of Accuracy and Precision 

By iteratively adding and removing different 

data sets from the ascending tables, MacDonald 

(1992) determined that a minimum of 40 data sets 

were needed to develop consistent and reliable 

guidelines. Clearly, some variability in the guidelines is to be expected as data are added or deieted, but, once the minimum amount of data is 

compiled, this variability appears to be minimal. 

MacDonald (1992) generally doubled or 

tripled the amount of data in the ascending tables 

compiled by Long and Morgan (1990) mainly 

with new data from field studies and laboratory 

spiked-sediment bioassays. Also, MacDonald 

(1992) considered only estuarine and marine data, 

thereby deleting the freshwater data included in 

Long and Morgan (1990). The effects on the 

guideline concentrations of eliminating some data 

and adding a substantial amount of new data are 

illustrated in Tables 14-1 and 14-2. The ERL and 

ERM values, based on the Long and Morgan 

(1990) data tables and the larger MacDonald 

(1992) tables, are compared by using the methods 

of Long and Morgan (1990) applied to both data 

sets. 

For 13 aromatic hydrocarbons, the average of 

the ratios between the two sets of guidelines was 

1.5 (1.9 for the ERb and 1.2 for the ERMs). For 

eight trace metals, the average of the ratios between the two sets of guidelines was 1.7. The 

trace metals ERL values changed more than the 

ERM values (average ratios of 1.9 for the ERLs 

and 1.5 for the ERMs). 

Overall, 7 of the 23 ERL values did not 

change and the ratios between the two sets of 

ERL values ranged from 1.0 to 9.4, Also, 7 of 

the 23 ERM values did not change. Of the 46’ 

values, 14 remained unchanged, 17 increased, and 

15 decreased. The overall mean factors of change 

were less than twofold for both trace metals and 

PAHs. These observations suggest that the guidelines are not terribly sensitive to the addition of 

new data once a minimum amount has been 

compiled. Also, they suggest that the guidelines 

originaliy developed by Long and Morgan (1990) 

generally are substantiated by additional data 

compiled by MacDonald (1992). 

The accuracy of the guidelines in predicting 

toxicity bas not ,yet been quantified. However, in 

the Hudson-Raritan estuary, the concentrations of 

many chemicals quantified in previous studies 

(Squibb et al., 1992) frequently exceeded the 

ERM guidelines in the Artbur Kill and rarely 

exceeded them in the lower Hudson River. In a 

recent survey funded by NOAA, sediments from 

the Arthur Kill were extremely toxic to amphipods 

and other species, whereas the sediments from the 

lower Hudson River were not toxic. 

14.4 STATUS 

14.4.1 Extent of Use 

The NSTP Approach is being used by 

NOAA’s National Status and Trends Program, by 

Environment Canada, and by the Florida Department of Environmental Regulation. A variation 

on the approach is being pursued by the California 

Water Resources Control Board. Other states and 

regional districts have inquired about the possible 

use of the approach. 

14.4.2 Extent to Which Appro8ch Has Been 

Field-Validated 

Validations of the guidelines have not yet 

b&n quantified. As described in Section 14.3.2.9, 

the original set of guidelines generaIly were 

substantiated by the addition of considerable 

amounts of new data, largely from field studies 

performed in many regions. The concordance 

between predictions of toxicity with the guidelines 

and actual observations of toxicity has been very 

14-14 
ZANSTP Appruuch 

Table 14-l. Ratios Between the Guideline Values for Polynuclear Aromatic Hydrocations Determined with 

Data from Long and Morgan (1990) and Those Determined with Data from MacDonald (lW2). 

Total number of data points available are listed (with those 

used to determine guidelines in parentheses). 

Chamiul MacDonald 

AnrlVtr ww 

Long 8nd 

bcw 

wm 

Ratlo BMwaon 

Two som 01 Valu# 

VdW 

heromod (+) 

Durusod (-) 

Pdynuclmrr womatlc hydrocarbon8 (fqb di) 

AtXIWphttlOM ~ww 

ERL 16 

ERM 500 

Anthracene n=88(46) 

ERL 65.3 

ERM 1100 

Fluorenne n&5(46) 

ERL 19 

ERM 540 

P-methylnathphalene n=49(28) 

ERL 70 

ERM 670 

naphthalens n=Q7(44) 

ERL 160 

ERM 2100 

phenanthrene n=l01(51) 

ERL 240 

ERM 1500 

bemzo(a)anthracene n=61(43) 

ERL 261 

ERM 1600 

benzo(a)pyrene n=69(44) 

ERL 430 

ERM 1600 

chysene n&9(45) 

ERL 364 

ERM 2600 

dibenzo(a,h)anthracen n=76(31) 

ERL 63.4 

ERM 260 

!brsntbns n=l17(71) 

ERL 600 

ERM 5100 

Pyreno n=QWJ 

ERL 665 

ERM 2600 

totaJ PAH n=76(34) 

ERL 4022 

ERM 44,760 

Mean change in PAH ERLs 

Mean change in PAH ERMs 

Overall mean change in PAH values 

r&%(15) 2.0(2.0) 

150 9.4 

650 1.3 

n=39(26) 23(1.6) 

85 1.0 l 

Q60 1.1 + 

n=44(28) 2.2(1.7) 

35 1.6 

640 1.2 . 

n&W (15) 1.6(1.Q) 

85 1.1 + 

670 1.0 . 

n=50(26) l.Q(l.6) 

340 2.1 

2laJ 1.0 . 

n=49(34) 2.1(1.5) 

225 1.1 + 

1380 1.1 + 

n=WW l.Q(l.4) 

230 1.1 + 

1600 1.0 . 

n=43 (27) 2.1(1.6) 

400 1.1 + 

25w 1.6 

n=41(27) 2.2(1.7) 

400 1.0 : 

2600 1.0 . 

n=23(18) 2.3(1.7) 

60 1.1 + 

260 . 

n=51(33) 

ki(l.8) 

600 1.0 . 

3600 1.4 + 

n43 (28) 2.2(1 xi) 

350 1.Q + 

2200 1.2 + 

n=WW 1.2(1 .O) 

4ow 1.0 . 

3%~ 1.3 + 

1.60 

1.17 

1.53 

II-15 
Sediment Classif~tion Methods Compendium 

Table 14-2. Ratios Between the Guideline Values for Total PCBs and Trece Metals Determined with 

Data from Long and Morgan (losO) and Those Detetmined with Data frwn MacDonald (1992). 

Total number of data points avaiiable are listed (with tiwse 

used to determine guidelines in parentheses). 

WPCB 

ERL 

ERM 

Trace YIlab (pfm dw.) 

n-126(50) *7WV l.S(l.5) 

22.7 50 22 - 

160 400 2.2 

w8mic n=143(227) 

ERL 6.2 

ERM 70.0 

cadmium f&61 (64) 

ERL 1.2 

ERM 9.6 

cww n=221(76) 

ERL 34.0 

ERM 270 

dwwnilun n=197(37) 

ERL 61 

ERM 370 

bd n=2f O(73) 

ERL 46.7 

ERM 223 

-w n-169(42) 

ERL 0.15 

ERM 0.71 

nidd n-169(lB) 

ERL 20.9 

ERM 51.6 

8ljker -=(=I 

ERL 1.0 

ERM 3.7 

zinc n=214(74) 

ERL 150 

ERM 410 

Meulctnng.hlPAHERL8 

MeanchmgahPAHERMs 

Overad mom change h ma& v&ma 

+WW 

33.0 

a.0 

n= 106(36) 

5.0 

9.6 

n-91 (51) 

70.0 

SBO 

~mw 

m 

145 

n&3(47) 

35.0 

110 

n-76(30) 

0.15 

1.3 

n=56(16) 

30 

50 

n=47(13) 

1.0 

2.2 

-ww 

120 

270 

1.74 

s.q1 .I) 

4.0 

1.2 

2.5(2.3) 

4.2 

:::(I .5) 

2.0 

::(I .6) 

1.0 

&.6) 

1.3 

2.0 

22(f -4) 

1.0 

1.5 

3.0(1 *l) 

1.4 

1.0 

2q1.9) 

1.0 

::;(I .6) 

1.25 

1.5 

1.9 

1.0 

- 

. 

. 

+ 

t 

+ 

. 

. 

. 

. 

+ 

+ 

+ 

< 

14-16 
IANSTP Approach 

good thus far, but the degree of concordance has 

not been quantified. Additional opportunities to 

field-validate the guidelines will be available in 

fulure studies in Tampa Bay, the Hudson-Raritan 

esluary, and southern California. 

14.43 Reasons for Limited Use 

The NSTP Approach initially was used by 

NOAA to develop informal guidelines for internal 

agency use. Therefore, knowledge of and access 

to the guidelines was limited. As interest in the 

guidelines increased, they were released in a 

government document with a limited distribution. 

Therefore, the main reason for the limited use of 

the approach has been the limited awareness of its 

existence. Furthermore, the equiitbrium-partitioning approach to national criteria and the most 

successful regional approach to criteria (apparent 

effects thresholds in Washington) have received 

considerable attention. Moreover, the guidelines 

thus far have not considered the potential for 

bioavailability or bioaccumulation because of a 

lack of data. 

14.4.4 Outlook for Future Use and Amount 

of Development Yet Needed 

There is significant potential for the expanded 

use of the NSTP Approach. Canada, Florida, and 

California currently are using the approach to 

develop their respective guidelines. Since the 

approach relies on existing data, other regionspecific guidelines could be developed easily, 

using the data available from specific regions. 

The approach can be used to validate criteria 

developed with other single-method approaches. 

The database can be accessed for specific regions 

or for fresh, estuarine, or marine waters. 

Several types of data are needed to further 

develop the approach. First, additional data are 

needed from studies in which TOC, grain size, 

and acid volatile sulfides were measured. Second, 

additional data are needed from spiked-sediment 

bioassays to establish cause-effect relationships. 

Third, additional data are needed from field 

studies in which very strong chemical gradients 

were observed. These studies should include 

measures of the toxicity and chemical contamination of bulk sediments and pore water. lhey 

would benefit from toxicity identification evaluations to identify the causative agents responsible 

for the observed biological effects (Ankley, 

1989). A number of large field surveys are under 

way and being planned by NOAA and will lead to 

additional data to be included in the database. 

Once these additional data are available, they 

could be entered into the database and used to 

develop updated or new guidelines. 

14.5 REFERENCES 

Adams, W.J., R. A. Kimerle, and J. W. Bamett, 

Jr. In press. Sediment quality and aquatic 

life assessment. Envir. Sci. and Technol. 

Haley, G. 1989. Sediment toxicity assessment 

through evaluation of the toxicily of interstitial water. Environmental Research Laboratory-Duluth. U.S. Environmental Protection 

Agency, Duluth, MN. 27 pp. 

Long, E. R. 1992. Ranges in chemical concentrations in sediments associated with adverse 

biological effects. Mar. Pollu. Bull. 24 (1): 

38-45. 

Long, E.R., D. MacDonald, and C Caimaoss. 

1992. Status and trends in toxicants and the 

potential for their biological effects in Tampa 

Bay, Florida. NOM Tech. Memo. NOS 

OMA 58. National Oceanic and Atmospheric 

Administration, Seattle, WA. 77 pp. 

Long, E.R., and R. Markel. 1992. An evaluation 

of the extent and magnitude of biological 

effects associated with chemical contaminants 

in San Francisco Bay, California. NOM 

Tech. Memo. NOS OMA 64. National Gceanic and Atmospheric Administration, Seattle, 

WA. 86 pp. 

Inng, E.R., and LG. Morgan. 1990. The potential for biological effects of sediment-sorbed 

contaminants tested in the National Status and 

Trends Program. NOAA Tech. Memo. NOS 

OMA 62. National Oceanic and Atmospheric 

Administration, Seatlle, WA. 175 pp. 

Lorenzato, S.G., A. J. Gunther, and J. M. 

O’Connor. 1991. Summary of a workshop 

14-17 
Sediment Class+ation Methods Compendium 

concerning sediment quality assesment and 

development of sediment quality objectives. 

California State Water Resources Control 

Board, Sacramento, CA. 32 pp. 

MacDonald, D.D. 1992. Development of an 

integrated approach to the assessment of 

sediment quality in Florida. Prepared for 

Florida Department of Environmental Regulation. MacDonald Environmental Services, 

Ltd. Ladysmith, British Columbia. 134 pp. 

MacDonald, D.D., and S.L Smith. 1991. A 

discussion paper on the derivation and use 

of Canadian sediment quality guidelines for 

the protection of freshwater and marine 

aquatic life. Prepared for Canadian Council 

of Ministers of the Environment. Environment Canada. Ottawa. 

MacDonald, D.D., S.L. Smith, M.P. Wong, and P. 

Mudroch. 1991. The development of Canadian marine environmental quality guidelines. 

Report prepared for the Interdepartment 

Working Group on Marine Environmental 

Quality Guidelines and the Canadian CounciI 

of Ministers of the Environment. Environment 

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