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OCR for page 100
A REVIEW OF THE DATA SUPPORTING THE
EQUILIBRIUM PARTITIONING APPROACH TO ESTABLISHING
SEDIMENT QUALITY CRITERIA
Dominic M. Di Toro
Manhattan College
The development of sediment quality criteria has been underway for
some time, and a number of approaches have been suggested (see Chapman,
1987 for a review). The discussion that follows presents the data and
interpretation that support the equilibrium partitioning approach (Pav-
lou and Weston, 1983; Pavlou, 1987) adopted for establishing sediment
quality criteria by the U.S. Environmental Protection Agency (EPA) Cri-
teria and Standards Division. The acknowledgment section of this paper
lists the many contributors to this effort. In this regard the author
should be viewed as a spokesman for workers involved.
Perhaps the first question to be answered is "why not use the exist-
ing procedure for the development of water quality criteria?" After
all, water quality criteria have demonstrated their utility; have been
reviewed by independent scientific groups; and--most important--a
methodology has been developed (Stephen et al., 1985) that presents the
supporting logic, establishes the minimum toxicological data set
required to develop a criteria, and specifies the numerical procedures
to calculate resulting criteria values. A natural extension would be
to apply these methods directly to sediments.
One reason water quality criteria have practical utility is that
they are based on straightforward measurements for most chemicals,
either total concentration or the recently proposed weak acid extrac-
tions for certain metals, and they appear to be broadly applicable.
The experience with site-specific modifications of the national water
quality criteria have demonstrated that the "water effect ratio" has
averaged 3.5 (Spehar and Carlson, 1984; Carlson et al., 1986~. The
implication is that subtle differences in water chemistry are not an
overwhelming impediment to nationally applicable, numerical water
quality criteria.
The primary impediment to direct application of the water quality
paradigm to sediment quality criteria is the use of total sediment
chemical concentration as a measure of bioavailable, or even poten-
tially bioavailable, concentration. This is not supported by the
available data (see, for example, Luoma, 1983~. A review of recent
~ ~ ~ ~~ . Different sediments can differ by
factors of ten or more in toxicity for the same total chemical concen-
tration of a toxicant. This is a severe obstacle since, without some
quantitative estimate of the bioevailable chemical concentration in a
sediment, it is impossible to evaluate its toxicity based on chemical
measurements.
~ _ ,
experiments is presented below
100
OCR for page 101
101
This is true regardless of the methodology used to assess biologi-
cal impact, be it field data sets comprising benthic biological and
chemical sampling or laboratory toxicity experiments. Without a unique
relationship between the chemical measurement and the biological end-
points, which applies across the range of sediment properties that
affect bioavailability, the cause and effect linkage is not support-
able. If the same total chemical concentration is ten times more toxic
in one sediment than another, how does one set a universal sediment
quality criteria that depends only on the total sediment chemical con-
centration? Some attempt must be made to address the issue of bioavail-
ability. Further it appears that any sediment quality criteria method-
ology that depends on chemical measurements in the sediment must face
this issue as well. It is not unique to the equilibrium partitioning
methodology.
BIOAVAILABILITY AND PORE WATER CONCENTRATION
The observation that provided the key ins ight to the solution of
the problem of quantifying the bioavailability of chemicals in sedi-
ments was that the dose-response curve for the biological effect of
concern could be correlated not to the total sediment chemical concen-
tration (pa chemical/g sediment) but to the interstitial water
(i.e., pore water) concentration (pa chemical/liter pore water).
Since this observation is a critical part of the logic behind the
equilibrium partitioning approach to sediment criteria, a substantial
amount of data has been assembled to support this observation. The
data are presented in a uniform fashion in Figures 1 through 5. The
biological response variable--survival rate, growth rate, body
burden--is plotted versus the total sediment concentration in the top
panel and versus the measured pore water concentration in the bottom
panel.
The kepone experiments, Figure 1 , are particularly dramatic (Adams
et al., 1985; Ziegenfuss et al., 1986~. Consider first the top panels .
For the low organic carbon sediment (0.09 percent) the fiftieth per-
centile total kep one concentration for both C. tentans mortality
LC50 and growth rate reduction EC50 are < 1 Agog. By contrast
the 1.S percent organic carbon sediment EC50 and LC50 are approxi-
mately 8 and 10 Agog respectively. The high organic carbon sedi-
ment (12 percent) has still higher LC50 and EC50s on a total
sediment kepone concentration basis (42 and 49 ~g/g). However, as
shown in the bottom panels, essentially all the data collapse into a
single curve when the pore water concentrations are used as the corre-
lating concentrations. Possible reasons for this observation will be
discussed below. It is important at this stage only to realize that on
a pore water basis the biological responses are essentially the same
for the three different sediments: the EC50 = 23 ~g/liter and LC50
= 28 ~g/liter, whereas when they are evaluated on a total sediment
kepone basis they exhibit an almost 50-fold range in kepone toxicity.
Figure 2 presents similar data for fluoranthene and cadmium and the
marine amphipod Rhepoxynius (Kemp and Swartz, 1986; Swartz et al.,
OCR for page 102
102
1987~. The results of the fluoranthene experiments parallel those for
kepone. The lowest organic carbon fraction sediments, 0.2 percent exhi-
bits the lowest LC50 on a total sediment concentration basis (3.1
g/g) and as the organic carbon concentration increases the LC50s
increase (6.7 and 11 ~g/g). On a pore water basis, however, the
data collapse to a single dose response curve. The cadmium experiments
were done using constant pore water concentrations and a sediment
amended with varying quantities of organic carbon. The unamended and
0.25 percent additional organic carbon exhibit essentially similar
responses. However the 1 and 2 percent amended sediments had much
higher LC50s. Using the pore water concentrations again collapses
the data into one dose response curve.
Figure 3 presents data for DOT and endrin and the freshwater
amphipod Hyal el la (Nebeker and Schuytema, 1988~. The response is
again similar to that observed above. On a total sediment
concentration basis, the organism response is different for the
different sediments. On a pore water basis, however, the dose
responses are similar.
a.
.
b.
80L
a\
X ORGANIC CARBON
.09
1.5
12.0
~ 60
_4
Or
~ 4 h ~
ae 21 ~ \ \ 1
0 old en n 40 . 0 60 . O ec . O
100
80 \
> 60 \
_ \
¢ \
~ 40 · \
2 1~
L
o
\
20 0 40.0 60.u
SEDIMENT KEPONE (ug/g)
50 100 so
PORE WATER KEPONE (ug/L)
I_
_ 12!
lo
7!
, i
X ORGANIC CARBON
.09 ~
1.5
12 0
~ ~ - -I
\ · \
-
a ~ ~ 0 40.0 60
SEDIMENT KEPONE (ug/g
at;
_ . _
0 25 50 75 100
PORE WATER KEPONE (ug/L)
FIGURE 1 Acute (a) and chronic (b) toxicity of kepone to Chironomus
tentans . SOURCE: Adams et al., 1983 e
OCR for page 103
a.
Bt
_I
~ 6(
rr
(n 4t
~_
r,
10`
~ ~ I I I
80 \
.. ] · \
~ 6 _ ~ -
L' ~
U) 40 _ }$
n
20.0
J
~o o
~-- J0
PORE WATER FLUORANTHENE (ug/L)
b.
,°O1
% ORGANIC CARBON
0.2 -
0.3 ~ -
0.5 ~
Rn
> 60
>
\~ _ ~ 40
~ \ ~
~ ~ —I ~
0 0 5 0 10 0 15 0
SEDIMENT FLUORANTHENE (ug/g)
\ \
l
~,j,,,
\ \
, ._ ,
% ORGANIC CARBON
Unamended -
+0.25 ~ -
+~.00
+2 00
\
~ollL
81
> 61
>
Lll 44
~e
~ 0 0 ~ 0 30 0 ~0 0 50 0 60 0 70 0
SEDIMENT CADMIUM (ug/g)
\
\
\
~ , __ _ 1- -;
0 1000 POOO 3000 4000 5c_
PORE Wt~lER CADMIUM (ug,'l.)
FIGURE 2 Acute toxicity of fluoranthene (a) and cadmium (b) to
Rhepoxynius abronius . SOURCE: Swartz et al., 1987.
a.
b.
.0. ,
so 1 \\
c ~ _t, ~ ~ , _ . , ~
O tOO t" ~0
SEDIMENT ODT (ug/g)
l
ORGANIC CAQBOb
3.0 ~
7.2 ~ -
10 5 ~
~no
~a
O. , _ ., ~ _ ~ o
0.00 1.00 2.00 3.00 4.00 5.00 0.0
PORE WATER DDT (ug/L)
l
% OQGANIC CARBON
3.0
6.1
11.2
ol - - _
o.o ~o.o 20 0
SEDIMENT ENDRIN (ug/g)
.
40 ~`
2 ~
2.0 4.0 6.0 ~ 0 .0
PORE WATER ENDRIN (ug/L)
FIGURE 3 Acute toxicity of DDT (a) and endrin (b) to Hyal el la .
SOURCE: Nebeker and Schuytema, 1988.
OCR for page 104
a.
~nnn ~ , ~ ~'~lil I I ' ' ""'
100
tD
a.
J
a~
b.
/ % ORGAN I C C
,? < 0 1 -
2
7
~ ~ ~ I I Itll ~ I I I ~ 1111 ~ I I I I I III I
,0 100 1000
SEO I MENT PERME rMn I N (U9 /9 )
~ooo~
100
c
_
0
m
C)
o
a~
T''
r.
C.
a
>
~_
,,,,,,
1 0000 1
z
I1J
o
m
~ 1C
o
m
;/ % ORGAN I C C
2
~ 7
l l l lull
10 100 1000
SEDIMENT CYPERMETHRIN (U9/9)
ennnn ~ ~', ~ ~ ~ ','',~ I I I Ililil l l "~''!
1111 1 1 1 11111' I r I '11111 I T ',r,~
~ ~ 0,~oooF
· 3: ~ Z
~oo
0 ~ I I I Illl
0 ol 0 1
1,,, , ~
1 to ~Oo
PORE WA T E R PE RMF ~ HR I N lug /L ~
z
CE
~ ~o - AW
-.w
ILi'~ll''l'.~""~'''""'~
0 0 1 0 ~ ~ ~o 100
PORE WATER CYPERMETMRIN (ug/L}
FIGURE 4 Bioaccumulation of permethrin (a) and cypermethrin (b) in
Chironomus tentans. SOURCE: Muir et al., 1985.
,00F
80r
6`
100 I 11111111_ I I ~' · IIIlIIII I I Irlll
.~\
so \ _
40 _
20 _
O I I 1 """ 1 ' """' .l ~ '111"' ~ ~ 0111. 1 1 .111
0° 103 102 )0 1 10 tOs
SEO I ME N ~ C&~ 1 ~ (ug /9 )
·:
~nr~
~_
60 _
40 _
20 _
b.
> 60
>
u, 40
X
20
C I
- ~ _ _
I 111141t 1 1 1: 1.1111111 1, 1 1111111 ~ 1 1~111
|0O 10 ~ ~o2 lo3 lo. 105
SEDIMENT CADMIUM lug/g)
tOOr—~T'
~-
hATER ONLY 0 \
SEDIMENT EXPOSURE · \
0 I,.~.,.. , .,,,,.s , .~.,.~s l,,,,... ,,,.,lm ,~ .~ - ..~....
10 ~o 3 lo~2 10-1 10° 101 ~o2 ~o3 lo4
Cd2+ RORE WATER (mg/L)
gOt
80t
J 70
> 60
_4
> 50
:) 40
at 30
20
~o
I I 111115 I Illrrr' I ~ 111118 1 1 Illror I I Illlm' I Illlrm r I 1111
0~
y
: , ,lill~ l Ill~lle , llllm l IlelIm l l=~- l llll" ~ lull
4 ~o ~3 ~o~2 '0- 1 ~Oo ~ lo2 to3 t04
CADMIUM PORE WATER (mg/L)
FIGURE 5 Acute toxicity of cadmium to Ampelisca (a) and Rbe-
poxynius (b). SOURCE: Scott and DiToro, 1988; Swartz et al., 1985.
OCR for page 105
105
Data for another biological endpoint, namely organism body burden,
are examined in Figure 4. Two synthetic pyrethroids, cypermethrin and
permethrin, and C. tentans were used (Muir et al., 1985~. Three
sediments, one of which was laboratory grade sand, were employed. The
bioaccumulation from the sand was approximately an order of magnitude
higher than the organic carbon containing sediment for both cypermeth-
rin and permethrin (top panels). On a pore water basis, however, the
bioaccumulation appeared to be linear (the lines are slope = 1) and
independent of sediment type (bottom panels).
Two sets of data are compared in Figure 5. which make an additional
point. They compare the response of Rhepoxynius (Swartz et al., 1985)
and Ampel isca (Scott and Di Toro, 1988) to cadmium in seawater-only
exposures and to measured pore water concentrations in sediment
exposures (lower panels). Note that the responses are the same with or
without the sediment present. The dose response curves using total
cadmium concentrations are also shown (top panels). It is interesting
to note that two organisms show essentially the same sensitivity to
cadmium. Yet the total cadmium LC50s differ by almost two orders of
magnitude (25 and 2000 Agog respectively) for the different
sediments.
These observations--that organism dose response curves for differ-
ent sediments can be collapsed into one curve if pore water is consi-
dered as the dose concentration--can be interpreted in a number of
ways. However, from a purely empirical point of view it suggests that
if it were possible to either measure the pore water concentration of a
chemical, or to predict it from the total sediment concentration and
the relevant sediment properties, then that concentration could be used
to quantify the chemical dose the sediment would deliver to an organ-
ism. Thus one is lead to examine the state of the art with respect to
predicting the partitioning of chemicals between the solid and liquid
phase. This is examined in the next section.
PARTITIONING OF CHEMICALS
A discussion of modeling sorption to particles is best organized by
classes of chemicals. For nonpolar hydrophobic organic chemicals sorb-
ing to natural soils and sediment particles, a number of empirical
models have been suggested (see Karickhoff, 1984 for an excellent
review). The characteristic that indexes the hydrophobicity of the
chemical is the octanol-water partition coefficient, KoW. The impor-
tant particle property is the mass fraction of organic carbon, foe.
For particles with foc ~ 0~5 percent the organic carbon appears to be
the predominate sorption phase. The only other important environmental
variable appears to be the particle concentration itself (O'Connor and
Connolly, 1980~. For the reversible (or labile) component of sorption,
a model has been proposed that predicts the partition coefficient of
nonpolar hydrophobic chemicals over a range of nearly seven orders of
magnitude with a logic standard error of 0.38 (Di Toro, 1985~.
For this class of chemicals the partitioning problem appears to be
solved. If cpOre is the aqueous pore water concentration (pa chemi-
cal/liter pore water), and r is the solid-phase concentration (pa
OCR for page 106
106
chemical/g sediment), then defining the partition coefficient ~ as
r ~ ~ Cpore'
then for sediment-pore water partitioning, ~ is given almost exactly
by1
~ focKow
(2)
What is important about this equation is that the partition coefficient
for this class of chemicals is linear in the organic carbon fraction
foe. As a consequence, the relationship between solid-phase
concentration r and pore water concentration cpOre can be written
r = fOcKowcpore'
or:
r = KOwcpore.
oc
If we define
rOc E
foc
as the organic carbon normalized sediment concentration (pa
chemical/g organic carbon) then
rOC ~ KOwcpore-
(3)
(4)
(5)
(6)
Hence we arrive at the following conclusion: for a specific chemical
with fixed how the organic carbon normalized total sediment concen-
tration rOC is proportional to the pore water concentration cpOre.
ORGANIC CARBON NORMALIZATION
From the above discussion we conclude that if a dose-response curve
correlates to pore water concentration, it should correlate equally
well to organic carbon-normalized total chemical concentration indepen-
dent of sediment properties. Of course this only applies to nonpolar
hydrophobic organic chemicals, since the rationale is based on a parti-
tioning theory for these chemicals. Figures 6 through 8 present these
comparisons. The lower panels present the response-total sediment
concentration data which is organic carbon normalized (pa chemi-
cal/g organic carbon). The top panels repeat the response-pore water
concentration plots. Note that in all cases the correlation is
~ The exact equation is logic ~ ~ logic foc + 0.0028 ~ 0.983
* logic KoW (Di Toro, 1985~.
OCR for page 107
~oo
l
1
Rn
a
20
O
o
I I I
— % ORGAN I C CARBON
·\ .09 -
.\ i.5 -
\ ~ 12 0 -
2'
50
,~
CD
Ct 25
25 50 75 ~ 00 ~ 25 ~ 50
PORE WATER KEPONE (ug/~)
~oo , , . , .
so \- -
c>r 60 ~ \j
Ln 40 ~ ~ \~
1
O tA ~
0 1 000 2000 :,(
SEDIMENT KEPONE (ug/g OC)
b.
50
125
J
o
tl
IOG
z
o
~ 75
~;
O _
o
15
% ORGAN I C CARBON
.09 ~
1.5
12.0
25 50 75
PORE WATER KEPONE lug/L)
25
C . _
0O 0 1000 2000
SEDIMENT KEPONE (ug/g OC)
FIGURE 6 Acute (a) and chronic (b) toxicity of kepone to Chironomus
tentans . SOURCE : Adams et al ., 1983 .
b.
300 ~ , ~ , __. ,
- , % OWGAN I C CARi]OI
~ ~ O
so \ 7.2
\ 10 5
. \
> 60 \
cr t
40 \
ae ~
20 \\
.
_ . _ _
o oo ~ oo ~ oo 3 oo ~ oo
PORE WATER ODT (ug/L)
> 60
>
a
n ~o
~e
10C. ~ , . . .
O" \
20
C
64
~4
CE
~7
X
. .~1 oL
0 500 tOOO 3500 0
SEDIMENT OD! (ug/g OC)
00 ~ ·
,~ % ORGAN I C CARBON
\ 3.0
AO ~ 6
\ 11 2
60 ~
·o \
20
~ v' _
5.00 0 0 2 0 4 0 6 0 ~ 0 1U O
PORE WAI ER ENORIN (ug/L)
~oa I} ' ' '
20
\
~-~
~ 00 200 300 400
SEDIMENT ENDRIN (ug/g OC)
FIGURE 7 Acute toxicity of DDT (a) and endrin (b) to Hyalella.
SOURCE: Nebeker and Schuytema, 1988.
OCR for page 108
108
FIGURE 8 Acute toxicity of
fluoranthene to Rhepoxynius
abronius. SOURCE: Swartz
et al., 1987.
10` . _~
10~.
I:
.
a\ ~
n
% ORIGAMI C CARSON
0.2 -
0.3 ~ -
0.5 ~
_ __
O G 20 0 40 0 [- 0 BO C
PORE WATER FLUORAN'THENE (ug/L)
l
it\,
~ 6G \ ~
> \
C \
U-J 40 \
20
.
. ~
n
0 3000 2000 3000 4000 5000 6000 7000
SEDIMENT FLUORANTHENE (ug/g OC)
essentially the same whether pore water or organic carbon-normalized
sediment concentrations are used for the chemical dose. This implies
that either of these concentrations can be used. Since it is much more
convenient to measure total sediment concentration and organic carbon
fraction the latter seems more practical.
EQUILIBRIUM PARTITIONING APPROACH TO SEDIMENT CRITERIA
The evidence presented above suggests that the pore water concen-
tration correlates to the biological responses examined. Hence the
biological effect levels generated in toxicity experiments can be
associated with the pore water concentrations. The procedure to set a
sediment quality criteria for a chemical would be to perform a series
of toxicity tests using benthic plants and animals, and establish the
range of effect concentrations based on the pore water concentrations.
The procedures set forth in the national guidelines (Stephen et al.,
1985) could be used directly. If it turned out that the most sensitive
benthic and pelagic species have similar sensitivity, then this would
be equivalent to requiring that the pore water concentration be at the
water quality criteria concentration. Hence the sediment quality
criteria, rSQc would be calculated using the water quality criteria
concentration CwQc, and the partition coefficient ~ , as follows:
rSQc = KpCWQC
(7)
OCR for page 109
109
where rSQc would be the total sediment concentration that is in equi-
librium with the pore water at the criteria level concentration. It is
from the equilibrium requirement that this approach is termed the "equi-
librium partitioning" method. With remarkable foresight this approach
was suggested for establishing sediment quality criteria by Pavlou and
Weston (1983) before the evidence discussed above was available. For
nonpolar hydrophobic organic chemicals, ~ ~ foc KoW, so that
rSQC = focKowCWQC'
(8)
Therefore, in order to compute a sediment quality criteria, rSQc, for
a particular chemical and sediment, one needs to know (1) the water
quality criteria, CwQc, (2) the octanol-water partition coefficient
of the chemical, KoW, and (3) the organic carbon fraction of the sedi-
ment,foc. A more easily remembered quantity is the organic carbon-
normalized sediment quality criteria rOC SQC, where
rOc SQC ~ _ e KowcwQc.
OC
This quantity is sediment independent, to the extent that organic car-
bon is the sole determinant of hydrophobic partitioning. It has been
suggested that the limit of applicability is foc > 0. 5 percent
(Karickhoff, 1984~. However, the fluoranthene data presented above sug-
gest that it might even apply to sediments with lower organic carbon
fractions. These procedures have been used to generate interim sedi-
ment criteria values (Cowan and Di Toro, 1988~.
THEORETICAL SPECULATIONS
The data presented above raise a number of interesting issues.
most surprising result is that the biological effects examined appear
to correlate to the interstitial water concentration, independent of
sediment type. This has been interpreted to mean that exposure is pri-
marily via the pore water, however the data correlate equally well to
the organic carbon-normalized sediment concentration. This may just
reflect the validity of the organic carbon-based, hydrophobic chemical
partitioning model.
However, another interpretation is possible. Consider the hypo-
thesis that the chemical potential (or fugacity) of a chemical controls
its biological activity. For a chemical dissolved in pore water at con-
centration cpOre, the chemical potential, Spore is
Spore ~ R0 + RTln(Cpore),
(10)
where pa is the standard state chemical potential, and RT is the
product of the universal gas constant and absolute temperature (Stumm
and Morgan, 1970~. For a chemical dissolved in organic carbon--
OCR for page 110
110
assuming that particle organic carbon can be characterized as a homo-
geneous phase--its chemical potential is
bloc ~ loo + RTln ~ rOc ),
(11)
where rOC is the weight fraction of chemical in organic carbon. If
the pore water is in equilibrium with the sediment organic carbon, then
Spore = Roc
(12)
The chemical potential that the organism experiences from either route
of exposure is the same. Hence, so long as the sediment is in equilib~
rium with pore water, the route of exposure is immaterial. Further if
chemical potential (or equivalently, fugacity) is proportional to biol-
ogical effects then the issue becomes "in which phase is ~ most
directly measured?"
Pore water concentration is an obvious suggestion. However, it is
necessary that chemical completed to colloidal--or as it is loosely
called, dissolved organic carbon (DOC)--be a small fraction of the
total measured concentration. If the partition coefficient for DOC is
on the order of KoW (Landr~,m, 1987), then the requirement is
CDOCKow << 1-
(13)
For logic KoW > 5 this may not be the case, and c ore would not
be a valid measure of the free chemical concentration.
Total sediment concentration normalized by sediment organic carbon
fraction is a second obvious choice. Note that this measurement is not
affected by DOC completing since that is affecting the distribution
within the aqueous phase but not the validity of Equation 11. The only
requirement is that sediment organic carbon be the only sediment phase
that contains significant amounts of the chemical. At foc < 0 5
percent this may no longer be the case.
SEDIMENT QUALITY CRITERIA FOR METALS
The equilibrium partitioning methodology for establishing sediment
quality criteria requires that effects concentration be determined in
an accessible phase and that the chemical potential of the chemical be
computed. The experiments presented above suggest that pore water con-
centrations of cadmium correlate to biological effect. A substantial
number of water column experiments point to the fact that biological
effects can be correlated to the divalent metal activity (Me +)
(Sunda and Guillard, 1976; Sunda et al., 1978; Anderson and Morel,
1978; Zamuda and Synda, 1982~. Hence the required partitioning model
should predict (Me +) in the pore water.
Models are available for cation and anion sorption to metal oxides
in laboratory systems (see the articles in Stumm [1987] for recent sum-
maries). The models for natural particles are less well developed.
Since the ability to predict partition coefficients is required if the
OCR for page 111
111
pore water metal concentrations are to be inferred from the total con-
centration, some practical model is necessary. -
A start in this direction was made during a recent conference (see
Di Toro et al., 19871. A more formal presentation is available (Jenne
et al., 1986~. The basic idea is that instead of only one site of sorp-
tion, organic carbon, as is assumed for nonpolar hydrophobic chemicals,
three sites of sorption are considered. In oxic soils and sediments
these have been identified as particulate organic carbon (POC) and the
oxides of iron and manganese (Jenne, 1968, 1977; Luoma and Bryan, 1981;
Oakley et al., 1982~. They are important because they have a large
sorptive capacity. Further they appear as coatings on the particles
and occlude the other mineral components. Thus they provide the pri-
mary sites for sorption of metals and they restrict the importance of
the clay and other mineral components of soils and sediments.
In addition to the sites of adsorption it is necessary to quantify
the fraction of total sediment metal that is chemically interacting
with the pore water. A substantial effort has been expended over the
years in attempting to determine the "bioavailable" portion of trace
metals in soils and sediments using chemical extractions (see Jenne
[1987] for a review and recommended procedure). The use of a relative-
ly mild reductant (hydroxylamine hydrochloride), which dissolves the Fe
and Mn oxides and liberates the sorbed metals, is recommended. The re-
ported results using extracted iron normalization (Tessier et al.,
1984) have been very encouraging. Similar results using an acid extrac-
tion have been found for arsenic in Nereis, a deposit feeding poly-
chaete, and Macoma, a deposit-feeding bivalve (Langston, 1980~; and
copper in aquatic plants (Campbell et al., 1985~. For mercury body bur-
dens in various benthic species (Langston, 1986), a strong correlation
exists between the sediment concentration normalized by organic matter
content.
The most direct evidence for the utility of the extraction-phase
normalization procedure, however, is from simultaneous observations of
interstitial water and sediment metal concentrations. Initial data of
this type (Tessier et al., 1985; Johnson, 1986) suggest that the extrac-
tion partitioning methodology can be used to establish metals criteria
in a way that directly addresses the bioavailability problem.
APPLICATION TO MARINE SEDIMENTS
The model presented above is directed at the oxic layer in fresh-
water sediments. It is not clear whether a similar model can be devel-
oped for estuarine and marine sediments. Certainly the role of metal
sulfides must be explicit in the formulation. Interestingly, some data
from estuaries (Langston, 1980, 1986) indicate iron normalization may
also apply. The application of the equilibrium partitioning approach
to these sediments is in the formative stage.
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112
ACKNOWLEDGMENTS
The EPA project director of the EPA Criteria and Standards Divi-
sion's sediment quality criteria effort is Christopher Zarba. The
organizations and personal involved are (in alphabetical order) the
Battelle Pacific Northwest Laboratory, Christina Cowan (Project
Leader); Everett Jenne; Drexel University, Herbert Allen; Enviro-
sphere, Spyros Pavlou; EPA ERL Corvallis, Richard Swartz; EPA ERL
Duluth, Nelson Thomas; EPA ERL Narragansett, David Hansen, John
Scott; Manhattan College, John Mahony. Many other people have
contributed ideas and data to the effort. The author particularly
thanks the scientists that provided raw data and results prior to
publication for inclusion in this review.
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Representative terms from entire chapter:
quality criteria
