1144

Environmental Toxicology and Chemistry, Vol. 18, No. 6, pp. 1144–1148, 1999q 1999 SETAC

Printed in the USA0730-7268/99 $9.00 1 .00

THREE-PHASE MODELING OF POLYCYCLIC AROMATIC HYDROCARBONASSOCIATION WITH PORE-WATER-DISSOLVED ORGANIC CARBON

SIDDHARTHA MITRA*† and REBECCA M. DICKHUT‡†Institute for Earth and Ecosystem Sciences, Department of EEOBiology, 310 Dinwiddie Hall, Tulane University,

New Orleans, Louisiana 70118, USA‡Department of Physical Sciences, School of Marine Science, College of William and Mary, Virginia Institute of Marine Science,

Gloucester Point, Virginia 23062, USA

(Received 11 November 1997; Accepted 26 October 1998)

Abstract—Log–log plots of measured organic carbon-normalized sediment pore-water distribution coefficients ( s) for severalK9OC

polycyclic aromatic hydrocarbons (PAHs) versus their octanol-water partition coefficients ( s) at two sites in the Elizabeth River,K9OW

Virginia, show large deviations from linearity. Organic-carbon normalized distribution coefficients for these PAHs between sedimentsand pore waters decreased by more than two orders of magnitude with depth as well. To determine to what extent pore waterdissolved and colloidal organic carbon (DOC) was responsible for the observed nonlinearity and decrease in , a three-phaseK9OC

model was used to estimate pore-water PAH-DOC binding coefficients (KDOC). Partitioning of PAHs to pore-water DOC (i.e., KDOC)enhances the observed ‘‘dissolved’’ phase PAH concentration, especially for high-KOW compounds, contributing to the nonlinearityin -KOW plots. However, our application of the three-phase partitioning model to these data indicate that, at most, pore-waterK9OC

PAH-DOC binding accounts for one order of magnitude of the observed decrease in with depth in the sediment bed. TheK9OC

results of this study are consistent with three-phase partitioning theory for hydrophobic organic compounds between sedimentorganic matter, pore-water DOC, and freely dissolved aqueous phases in natural systems.

Keywords—KDOC Polycyclic aromatic hydrocarbons Three-phase partitioning model Pore waterDissolved organic carbon

INTRODUCTION

In a two-phase aqueous system at equilibrium, the concen-tration of a hydrophobic organic contaminant (HOC), such asa polycyclic aromatic hydrocarbon (PAH), in the freely dis-solved phase (CW, mg/L) relative to that in a sorbed phase (CS,mg/kg) can be ideally depicted by a linear isotherm

KD 5 CS/CW (1)

where KD (L/kg) is the two-phase equilibrium distribution co-efficient. The KD describes the condition under which chemicalfugacities, or ‘‘escaping tendencies,’’ in the sorbed and freelydissolved phases are equal. Normalization to the amount ofsedimentary organic carbon (fOC) modifies Equation 1 to

KOC 5 KD/fOC (2)

where KOC is the organic carbon-normalized two-phase distri-bution coefficient. Equation 2 assumes that HOCs associatemainly with sedimentary organic carbon but incorrectly as-sumes that HOCs will associate identically with all types oforganic moieties. The KOC values for the same PAH are ob-served to vary by as much as a factor of 10 for differentsediment/soil types [1–5]. Thus, a general solution to this equa-tion cannot be used to predict HOC distributions to particulatematter a priori, as the association of HOCs by geosorbents andsubsequent distribution in aqueous systems can be quite com-plex and are not well understood [6].

* To whom correspondence may be addressed(smitra@mailhost.tcs.tulane.edu). The current address of S. Mitra isDepartment of EEOBiology, 310 Dinwiddie Hall, Tulane University,New Orleans, LA 70118, USA.

Contribution 2204 from the Virginia Institute of Marine Science.

In natural systems where organic carbon normalization doesnot account for observed differences in equilibrium distribu-tion coefficients, a third sorbing organic phase, or colloidalphase, has also been suggested to control HOC partitioning[7–9]. Colloids are especially important because of their highspecific surface area [10] and thus their potential adsorptivecapacity for HOCs in water. In estuarine sediments, wherecolloidal organic carbon in pore waters can contribute as muchas 90% of the measured dissolved organic carbon [8], HOCpartitioning to pore-water colloids is potentially an importantprocess affecting the contaminant’s distribution and bioavail-ability [9,11–15]. Three-phase partitioning of HOCs betweenparticles, colloids (high-molecular-weight DOC) and low-mo-lecular-weight DOC, and the freely dissolved phase is de-scribed by the following, adapted from [7]:

5 KOC·fOC/(1 1 KDOC·[DOC])K9D (3)

where is the observed or apparent distribution coefficientK9Dfor an HOC in a three-phase system in which colloid and DOC-associated HOCs are operationally defined as being in the dis-solved phase, KDOC is the colloid and DOC binding coefficient(L/mg organic carbon), and DOC is the concentration of col-loidal 1 dissolved organic matter (mg organic carbon/L). Nor-malizing in Equation 3 to fOC results in the apparent organicK9Dcarbon normalized distribution coefficient, herein referred to as

. From Equation 3, it is evident that as [DOC] or KDOCK9OC

decreases, the effect of colloidal or DOC binding is reduced.The affinity of a particular HOC for colloidal and dissolved

organic carbon depends in part on the hydrophobicity of thecompound as well as on the source of the natural organic matter[14]. Thus, the magnitude of the PAH-DOC binding coefficient

Modeling PAH-pore-water-dissolved organic carbon associations Environ. Toxicol. Chem. 18, 1999 1145

Fig. 1. Map of Elizabeth River and sampling sites.

(KDOC), in addition to the DOC concentration, determines theextent to which DOC complexation takes place [14]. The ob-jectives of this paper were to apply a three-phase partitioningmodel to observed sediment/pore-water PAH distributions andto determine (1) whether the observed nonlinearity in two-phase s of PAHs in the urban Elizabeth River estuary areK9OC

the result of PAH binding to a third dissolved phase and (2)to what extent PAH-pore-water s contribute to observedK9DOC

decreases in s with depth in the sediments.K9OC

METHODS

Sediment sampling

Sediment cores were taken from selected sites in the SouthernBranch of the Elizabeth River estuary, an urbanized tributary oflower Chesapeake Bay, Virginia, USA (Fig. 1). Preliminary sam-pling of sediments was conducted in July 1994 using a box core(26.5 cm 3 26.5 cm 3 60 cm). Sediments were subsampledusing (60 cm 3 7.5 cm i.d. 3 0.6 cm wall thickness) polyvinylchloride tubing, which was sealed to maintain anoxic conditions.Site 1 (368489000N, 768179390W) was ;100 m southwest of apublic boat launch in ;1.5 m of water. Site 2 (368489680N,768179410W) was located in about 1 m of water and situatedacross the main channel from a former wood treatment facilitythat historically was responsible for creosote spills to the area.On returning to the lab, sediments were extruded at 2-cm depthintervals and processed as described in the following discussion.Because of core compression and the sediment water content,only 22 cm of sediment were available to be processed at this

time. Samples were analyzed for particulate organic carbon (POC)in addition to pore-water and sediment-associated PAHs, as de-scribed in the following. No DOC samples were collected fromthese cores.

In a subsequent visit to each site in September 1995, ad-ditional box cores were collected as described elsewhere [16].Within 72 h, these subcores were extruded at 1-cm depth in-tervals. Samples were collected for POC, DOC, and PAH anal-yses as described elsewhere [16].

Separation of particulate and dissolved phases

For each depth interval, the 0.5-cm outer edge was dis-carded to minimize cross-contamination, and sediment wascentrifuged in pre-ashed (4008C for 4 h) pint-sized glass jarsat 1,500 g for 25 min. Overlying water was pipetted off andvacuum filtered through a 1-mm nominal-pore-size glass fiberfilter (combusted for 4 h at 4508C) to separate particulate andpore-water fractions. Thus, the pore-water filtrates containedboth freely dissolved and DOC-bound PAHs.

PAH and organic carbon analysis

Filtered pore water from each depth interval was extractedwith hexane (4 3 20 ml after addition of surrogate standards)and analyzed for PAHs as described in the following. Sedimentsamples (;5 g wet weight) were transferred to pre-ashed 50-ml tubes, to which a surrogate standard mixture containingfive deuterated PAHs (;1 mg PAH/g hexane) were added.Sediments were extracted with acetone (1 3 20 ml) and meth-ylene chloride (DCM) (4 3 20 ml) for all samples except thoseat Site 2 (September 1995), which were extracted with DCMonly (4 3 20 ml) by sonicating 1 h in an ultrasonic water bathand shaking for 2 min. The extracts were filtered through pre-cleaned Pasteur pipettes filled with solvent-rinsed glass wooland precleaned (Soxhlet extracted with DCM for 48 h) an-hydrous Na2SO4, rotoevaporated and solvent exchanged withhexane. The sample extract was then cleaned up by solid-liquidchromatography on silica to remove organic polymers, ali-phatic, and polar compounds [17]. Subsequently, the extractswere rotoevaporated, and an internal/recovery standard con-taining additional deuterated PAHs was added prior to volumereduction under a stream of N2. The PAHs were quantifiedrelative to the deuterated surrogate PAHs by gas chromatog-raphy/mass spectrometry using selected ion monitoring. Re-coveries ranged from 64.4 6 12.3% for naphthalene to 97.06 16.8% for less volatile PAHs.

Particulate organic carbon was quantified using a CHNS-Oanalyzer (Fisions, EA 1108, Beverly, MA, USA). Preweighedsediment samples were acidified with 6 N HCl and allowedto evaporate to dryness to remove inorganic carbon. The sam-ple was then placed in the analyzer and flash heated to 1,0208Cto convert organic carbon to CO2 [18]. Pore-water samples formeasurement of DOC were stored in pre-ashed vials, and theheadspace of each sample vial was purged with N2(g) prior tofreezing at 2808C. Dissolved organic carbon was subsequentlymeasured using a Shimadzu (Kyoto, Japan) TOC 5000 totalorganic carbon analyzer. Analysis was done by high-temper-ature oxidation of pore-water samples after acidification withone to three drops of 6 N HCl per 10 ml of sample [19].

RESULTS AND DISCUSSION

Preliminary results for sediments collected in July 1994 atSite 2 in the Elizabeth River indicated that PAH log sK9OC

deviated from linearity when plotted against log KOW (Fig. 2).

1146 Environ. Toxicol. Chem. 18, 1999 S. Mitra and R.M. Dickhut

Fig. 2. Relationship between and KOW for selected PAHs inK9OC

Elizabeth River sediment sampled at Site 2 in July 1994 (V: 0–2cm; M: 18–20 cm). Selected PAHs: naphthalene acenaphthylene,acenaphthene, fluorene, anthracene, phenanthrene, pyrene, fluor-anthene, benz[b]fluoranthene, chrysene, benz[a]anthracene, ben-zo[a]pyrene, benzo[e]pyrene, benz[k]fluoranthene, dibenz[ah]anthracene,benzo[ghi]perylene.

Table 1. Curve fit coefficients (6SE) for Equation 4 fitted to dataK9OC

for PAHs with log KOW # 5.78 (Site 1) and log KOW # 5.2 (Site 2)

Depthinterval(cm)

Site 1b 6 SEb

Site 2b 6 SEb

0–14–55–66–77–88–9

12–1314–1516–1719–2024–2526–2727–2830–3131–32

1.11 6 0.2011.35 6 0.185

1.37 6 0.164

1.19 6 0.1720.800 6 0.048

1.16 6 0.308

20.799 6 0.10720.587 6 0.50620.645 6 0.177

2.33 6 0.083

2.33 6 0.1611.88 6 0.090

2.16 6 0.182

2.16 6 0.378

2.06 6 0.1442.27 6 0.1211.94 6 0.622

2.71 6 0.179

Fig. 3. Relationship between and KOW for selected PAHs in Eliz-K9OC

abeth River sediment sampled in September 1995. Site 1, filled sym-bols (v: 0–1 cm; m: 17–18 cm); Site 2, open symbols (V: 0–1 cm;M: 19–20 cm).

Further, PAH distribution coefficients in these cores decreasedwith sediment depth despite normalizing to organic carbon.This can be seen by the spread of the data points betweendepth intervals (Fig. 2). However, pore water was not analyzedfor DOC in the preliminary 1994 experiment; thus, the impactof DOC on the observed ’s could not be evaluated. InK9OC

subsequent cores collected in September 1995, PAH ’sK9OC

measured at both sites also deviated from linearity when plot-ted against KOW (log–log plots) and at Site 1 exhibited as muchas a 2.5 order of magnitude decrease with depth (Fig. 3).Interestingly, the second Site 2 core showed little evidence ofa decrease in with depth (Fig. 3). The differences in mea-K9OC

sured ’s at Site 2 between 1994 and 1995 likely reflectK9OC

spatial heterogeneity of the sediments in this perturbed estu-arine environment. However, for the samples collected in 1995,

pore-water DOC was quantified so that a three-phase parti-tioning model (Eqn. 3) could be applied to the data.

Modeling of PAH-DOC binding

Because values of log for PAHs in Elizabeth RiverK9OC

sediments do not increase linearly with log KOW above alog KOW of about 5 to 6 (Figs. 2 and 3), we propose that athird phase for PAH binding in pore water is influencing thePAH distributions beyond log KOW ø 5.5. Assuming that lesshydrophobic PAHs bind minimally to DOC [14,20,21], a two-phase distribution coefficient model (linear free energy rela-tionship)

log KOC 5 a·log KOW 1 b (4)

was fit to the data for compounds with log KOW ø 5.5 or less.In Equation 4, a is the ratio of the change in free energy thatoccurs with PAH partitioning into natural organic matter tothe free energy change when PAHs associate with water-sat-urated octanol [22]. In contrast, b describes the relative affinityof the PAHs for natural organic matter compared to n-octanol.Theoretically, for a given set of compounds, a should equal 1and b should equal zero if PAHs partition to natural organicmatter in exactly the same manner as n-octanol. However,natural organic matter tends to be heterogenous in composi-tion. Thus, values of a and b in experimental -KOW plotsK9OC

diverge from theoretical approximations [1–3,13,23,24].The mean slope of the linear regressions between log K9OC

and log KOW data (Eqn. 4) was 0.89 6 0.37 (r2 5 0.78 6 0.21)for Site 1 (1995 data) and 0.84 6 0.41 (r2 5 0.80 6 0.26) forSite 2 (1995 data). The log KOW cutoff used for Site 1 was5.78, whereas log KOW 5 5.2 was used as the cutoff for fittingthe Site 2 data to the two-phase distribution coefficient model.Inclusion of the higher KOW compounds in the Site 2 data setsignificantly lowered (paired t test, p 5 .0002) values of a,indicating that the log values for PAHs with log KOW .K9OC

5.2 in this core are affected by DOC binding. For both sites,these slopes are not significantly different from a value of 1at the 95% confidence level. Thus, no conclusions can bedrawn about the impact of aromatic ring additions on the partialmolar excess free energies of PAHs in natural organic matterversus n-octanol [22].

Values of b were also determined from the linear regres-sions of log -log KOW (Eqn. 4) for each depth intervalK9OC

Modeling PAH-pore-water-dissolved organic carbon associations Environ. Toxicol. Chem. 18, 1999 1147

Table 2. Variables and curve fit results for three-phase partitioning model (Eqn. 5) fits to Elizabeth River core data

Depth (cm)

Site 1

fOC

DOC(mg/L) l 6 SEl

Site 2

fOC DOC l 6 SEl

0–12–3a

4–5

0.01630.01630.0137

13.99.939.89

26.46 6 0.17126.52 6 0.12726.52 6 0.316

0.0403 7.39 26.16 6 0.070

6–719–2024–25

0.02370.03110.0179

8.9731.510.32

26.18 6 0.10726.72 6 0.08426.37 6 0.067

27–2830–31

0.0222 21.6 26.69 6 0.2470.0187 9.82 26.06 6 1.31

a b interpolated between 0–1 cm and 4–5 cm.

Fig. 5. Modeled pore-water DOC binding coefficients, KDOC (l·KOW).Site 1, filled symbols; Site 2, open symbols.

Fig. 4. Measured (symbols) and predicted (lines) log values forK9OC

PAHs in Elizabeth River sediments assuming DOC binding of PAHswith log KOW . 6.

(Table 1). For the cores collected in 1995, values of b at Site2 were higher than those at Site 1, indicating a greater affinityof PAHs for the sediment organic carbon at Site 2 comparedto Site 1. The fact that log s fall consistently above the 1:K9OC

1 line when plotted against log KOW at Site 2 for the 1995samples (Fig. 3) illustrates that the sediments in this core area more attractive substrate for PAHs than n-octanol (i.e., b .0). In contrast, Site 1 sediments from 1995 and Site 2 sedimentsfrom 1994 appear to be a less desirable substrate for PAHs,as s fall near or below the 1:1 line (Figs. 2 and 3). ValuesK9OC

of b for the Site 1 data are lower overall compared to Site 2(1995 data) and in some cases less than zero (Table 1). Thebimodal distribution in down-core values of b at Site 1 (Table1) is likely a measure of the difference in PAH accessibilityto sedimentary organic matter at this site. The higher valuesof b in the top portion of the core might be due to the presenceof more accessible sedimentary organic matter for PAH bind-ing, whereas the negative values of b near the bottom of thecore from Site 1 coincide with less accessible particulate or-ganic matter [16].

To determine PAH binding coefficients to DOC, nonlinearcurve fitting of the log-transformed version of Equation 3

log K 9 5 (a · log K 1 b) 1 log fD OW OC

logK 1lOW2 log(1 1 [DOC] · 10 ) (5)

was used, substituting . This substitution islog K 1lOW10 for KDOC

derived from the linear free energy relationship describing

PAH partitioning to dissolved and colloidal organic carbon inproportion to KOW

log KDOC 5 a·log KOW 1 l (6)

and is analogous to Equation 4 for particulate organic matter/water partitioning, assuming that a 5 1. From Equation 5, itis apparent that when log KOW, l, and [DOC] are small, log(11 [DOC]· ) approaches zero and /fOC 5 KOC. InlogK 1lOW10 K9Dcontrast, when log KOW, l, and [DOC] are large, the measuredpartition coefficient becomes much lower than expected, em-phasizing the importance of colloid binding for more hydro-phobic compounds. For our Elizabeth River (1995) sedimentcores, l in Equation 5, which describes the relative affinity ofPAHs for DOC, is the only unknown; all other variables wereeither measured ( , fOC, DOC, KOW) or estimated (KOC). Val-K9Dues of l determined from Equation 5 (Table 2) had relativestandard errors averaging 4.5 (6 6.9%).

The three-phase partitioning model, along with the as-sumptions that log KOC and log KDOC are linearly related tolog KOW, describe the observed distribution coefficients well(Fig. 4). Note, however, that l varies only by ;1.25% betweenboth sites (Table 2), indicating that KDOC varies by ;20% (i.e.,

1148 Environ. Toxicol. Chem. 18, 1999 S. Mitra and R.M. Dickhut

less than one order of magnitude with depth) (Fig. 5). Thus,DOC binding of pore-water PAHs is responsible for less than1 log unit of the observed 1- to 2.5-log-unit decrease in K9OC

with depth in the sediment.In summary, PAH partitioning to pore-water DOC does

account for less-than-predicted values of for high-molec-K9OC

ular-weight PAHs. However, PAH binding to pore-water DOCexplains only a fraction of the observed decreases in withK9OC

depth in sediments from the urban Elizabeth River estuary.Most notably, our results indicate that PAH binding to pore-water DOC is consistent across two sites with widely differentsediment geochemistry and deposition environments [16].

Acknowledgement—The authors wish to thank Pat Calautti, Tim Del-lapena, Ginger Edgecombe, Kurt Gustafson, Kimani Kimbrough,Kewen Liu, Elizabeth MacDonald, Linda Meneghini, Krisa Murray,Linda Schaffner, Michelle Thompson, and Andy Zimmerman for theirsupport during the project. Also thanks to the VIMS Analytical Ser-vices Lab and to the captain and first mate of the R/V Bay Eagle.This research was sponsored by Office of Naval Research HarborProcesses Program Grant N00014-93-1-0986 and N00014-96-1-0062.

REFERENCES

1. Weber WJ Jr, Voice TC, Pirbazari M, Hunt GE, Ulanoff DM.1983. Sorption of hydrophobic organic contaminants by sedi-ments, soils and suspended solids—II: Sorbent evaluation studies.Water Res 17:1443–1452.

2. Garbarini DR, Lion L. 1986. Influence of the nature of soil or-ganics on the sorption of toluene and trichloroethylene. EnvironSci Technol 20:1263–1269.

3. Gauthier TD, Seitz WR, Grant CL. 1987. Effect of structural andcompositional variations of dissolved humic materials on pyreneKOC values. Environ Sci Technol 21:243–247.

4. Gratwohl P. 1990. Influence of organic matter from soils andsediments from various origins on the sorption of some chlori-nated aliphatic hydrocarbons: Implications on KOC correlations.Environ Sci Technol 24:1687–1693.

5. Weber WJ Jr, McGinley PM, Katz LE. 1992. A distributed re-activity model for sorption by soils and sediments. 1. Conceptualbasis and equilibrium assessments. Environ Sci Technol 6:1955–1962.

6. Luthy RG, Aiken GR, Brusseau ML, Cunningham SD, GschwendPM, Pignatello JJ, Reinhard M, Traina SJ, Weber WJ Jr, WestallJ. 1997. Sequestration of hydrophobic organic contaminants bygeosorbents. Environ Sci Technol 31:3341–3347.

7. Morel FMM, Gschwend PM. 1987. The role of colloids in thepartitioning of solutes in natural waters. In Stumm W, ed, AquaticSurface Chemistry. John Wiley & Sons, New York, NY, USA,pp 405–422.

8. Burgess RM, McKinney RA, Brown WA, Quinn JG. 1996. Iso-lation of marine sediment colloids and associated polychlorinated

biphenyls: An evaluation of ultrafiltration and reverse-phase chro-matography. Environ Sci Technol 30:1923–1932.

9. Burgess RM, McKinney RA, Brown WA. 1996b. Enrichment ofmarine sedimentary colloids with PCBs: Trends resulting fromPCB solubility and chlorination. Environ Sci Technol 30:2556–2566.

10. Hiemenz PC. 1986. Principles of Colloid and Surface Chemistry.Marcel Dekker, New York, NY, USA.

11. McCarthy JF, Jimenez BD. 1985. Reduction in bioavailability tobluegills of PAHs bound to dissolved humic materials. EnvironToxicol Chem 4:511–521.

12. Landrum PF, Nihart SR, Eadie BJ, Herche LR. 1987. Reductionin bioavailability of organic contaminants to the amphipod, Pon-toporeia hoyi, by dissolved organic matter of sediment interstitialwaters. Environ Toxicol Chem 6:11–20.

13. McCarthy JF. 1989. Bioavailability and toxicity of metals andhydrophobic organic contaminants. In MacCarthy P, Suffett IH,eds, Aquatic Humic Substances—Influences on Fate and Trans-port of Pollutants. Advances in Chemistry Series 219. AmericanChemical Society, Washington, DC, pp 263–277.

14. Di Toro DM, et al. 1991. Technical basis for establishing sedimentquality criteria for nonionic organic chemicals using equilibriumpartitioning. Environ Toxicol Chem 10:1541–1583.

15. Martin JM, Dai MH, Cauwet G. 1995. Significance of colloidsin the biogeochemical cycling of organic carbon and trace metalsin a coastal environment—An example of the Venice Lagoon(Italy). Limnol Oceanogr 40:119–131.

16. Mitra S. 1997. PAH distributions within urban estuarine sedi-ments. PhD thesis. College of William and Mary, GloucesterPoint, VA, USA.

17. Dickhut RM, Gustafson KE. 1995. Atmospheric inputs of selectedpolycyclic aromatic hydrocarbons and polychlorinated biphenylsto southern Chesapeake Bay. Mar Pollut Bull 30:385–390.

18. Verado DJ, Froelich PN, McIntyre A. 1990. Determination oforganic carbon and nitrogen in marine sediments using the CarloErba NA-1500 Analyzer. Deep-Sea Res 37:157–165.

19. Sugimura Y, Suzuki Y. 1988. A high-temperature catalytic oxi-dation method for the determination of non-volatile dissolvedorganic carbon in seawater by direct injection of a liquid sample.Mar Chem 24:105–131.

20. Eadie BJ, Morehead NR, Landrum PF. 1990. Three-phase parti-tioning of hydrophobic organic compounds in Great Lakes Wa-ters. Chemosphere 20:161–178.

21. Achman DR, Brownawell BJ, Zhang L. 1996. Exchange of po-lychlorinated biphenyls between sediment and water in the Hud-son River estuary. Estuaries 19:950–965.

22. Schwarzenbach R, Gschwend PM, Imboden DM. 1993. Environ-mental Organic Chemistry. John Wiley & Sons, New York, NY,USA, pp 134–137.

23. Karickhoff SW. 1984. Organic pollutant sorption in aquatic sys-tems. J Hydraul Eng 110:707–735.

24. Huang W, Young TM, Schlautman MA, Yu H, Weber WJ Jr. 1997.A distributed reactivity model for sorption by soils and sediments.9. General isotherm nonlinearity and applicability of the dualreactive domain model. Environ Sci Technol 31:1703–1710.