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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

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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.,

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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

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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.

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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.

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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

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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~.

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~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.

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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)

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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--

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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

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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. REFERENCES Adams, W. J., Kimerle, R. A. and Masher, R. G. 1985. Aquatic safety assessment of chemicals sorbed to sediments. In Aquatic Toxicology and Hazard Assessment: Seventh Symposium, R. D. Cardwell, R. Purdy and R. C. Bahner, eds. Philadelphia: American Society for Testing and Materials. Pp..429-453. Anderson, D. M. and Morel, F. M. M. 1978. Copper sensitivity of Gonyaulax camarensis. Limnol. Oceanogr. 23:283-29S. Campbell, P. G. C., A. Tessier, M. Bisson, and R. Bougie, 1985. Accumulation of copper and zinc in the yellow lily, Nuphar variegatum: relationships to metal partitioning in the adjacent lake sediments. Can. J. Fish. Aquat. Sci. 42:23-32. Carlson, A. R., H. Nelson, and D. Hammermeister. 1986. Development and validation of site-specific water quality criteria for copper. Environ. Toxicol. and Chem. 5:997-1012. Chapman, G. A. 1987. Establishing sediment criteria for chemicals-- regulatory perspective. In Fate and Effects of Sediment--Bound Chemicals in Aquatic Systems, K. L. Dickson, A. W. Maki, and W. A. Brungs, eds. New York: Pergamon Pres. Pp. 35S-376. Cowan, C. E. and D. M. Di Toro. 1988. Interim Sediment Criteria Values for Nonpolar Hydrophobic Compounds. Richland, Washington: Battelle Pacific Northwest Laboratories. Di Toro, D. M. 1985. A particle interaction model of reversible organic chemical sorption. Chemosphere 14~10~:1503-1538. Di Toro, D. M., F. Harrison, E. Jenne, S. Karickhoff, and W. Lick. 1987. Synopsis of discussion session 2: Environmental fate and compartmentalization. In Fate and Effects of Sediment-Bound Chemicals in Aquatic Systems, K. L. Dickson, A. W. Maki, and W. A. Brungs, eds. New York: Pergamon Press. Pp. 136-147. Jenne, E. A. 1968. Controls on Mn, Fe, Co, Ni, Cu. and Zn concentra- tions in soils and water--the significant role of hydrous Mn and Fe oxides. In Advances in Chemistry. Washington, D.C.: American Chemical Society. Pp. 337-387. Jenne, E. A. 1977. Trace element sorption by sediments and soil--sites and processes. In Symposium on Molybdenum in the Environment9 Vol.

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