Sorption behaviour of some chlorophenols in lake aquatic humic matter

Talanta 56 (2002) 523– 538 www.elsevier.com/locate/talanta

Sorption behaviour of some chlorophenols in lake aquatic humic matter Juhani Peuravuori a,*, Nina Paaso b, Kalevi Pihlaja a a

Department of Chemistry, Physical Chemistry, Uni6ersity of Turku, FIN-20014 Turku, Finland b Electric power company of TVO (Teollisuuden Voima Oy), FIN-27160 Olkiluoto, Finland

Received 29 June 2001; received in revised form 21 September 2001; accepted 3 October 2001

Abstract The sorption behaviour of 2,4-dichlorophenol (2,4-DCP), 2,4,6-trichlorophenol (2,4,6-TCP), 2,3,4,5-tetrachlorophenol (2,3,4,5-TeCP) and pentachlorophenol (PCP) with an aquatic humic sorbent (HS) was examined in their single and mixed solutions at different acidities (pH 3, 5.5 and 7). The binding capacities and equilibrium coefficients (KOC) obtained were fairly close to the literature values but still underline HS’s structural and steric influence on the sorption. The most acidic carboxylic (COOH) groups of the HS structure have unquestionably an essential role in the sorption. The amounts of different chlorophenols bound onto the constant quantity of the aquatic HS were in reality very low demonstrating that the amount of the dissolved organic carbon (DOC) in the environment plays a greater role than the value of KOC. The ability of the aqueous phase to force chlorophenols to associate with the HS becomes at more neutral acidities weaker and weaker and other binding mechanisms become favoured in comparison to hydrogen or hydrophobic bonds. Sorption isotherms were constructed from sorption data, and conformity to a linear model, non-linear Freundlich equation and Langmuir equation was checked. © 2002 Elsevier Science B.V. All rights reserved. Keywords: Sorption coefficients; Chlorophenols; LFER; Humic substances

1. Introduction All chlorophenols are practically solids (melting points from ca. 33 to 191 °C). Their octanol – water partition coefficients (KOW) strongly increase with the number of chlorine atoms and water solubilities (hydrophilicity) inversely decrease. Also, the degree of dissociation of chlorophenols * Corresponding author. Tel.: +358-2-333-6757; fax: + 358-2-333-6700. E-mail address: [email protected] (J. Peuravuori).

increases (indicated as descending pKa values) with increasing number of chlorine atoms. Chlorophenols form a class of polarizable hydrophobic compounds. Although the use and utilization of chlorophenols have decreased during recent years, they still cause an environmental problem at many locations, e.g. as contaminants of sediments and even ground-waters [1–3]. The understanding of the behaviour of chlorophenols requires an assessment of the processes influencing their environmental fate, transport and bioavailability in soils and natural waters. The

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study of interactions of organic contaminants with the natural organic matter is also of great importance for the molecular chemistry. Much of the research on sorption of chlorophenols is based on equilibrium measurements. On the other hand, it has been stressed [4 –6] the importance of the influence of sorption non-equilibrium on organic contaminant transport. Non-equilibrium sorption has been further divided into three categories: physical and chemical non-equilibrium and intrasorbent diffusion (intraorganic matter or intramineral diffusion). Different sorts of mathematically quite complex approaches have been applied to simulate sorption kinetics, e.g. a biocontinuum model has been very popular. In the acidity (pH) range relevant to most environmental scenarios some of the more chlorinated phenols can become partially or almost totally ionised making the sorption – desorption behaviour more and more complicated. Inherent to chemistry is the concept that a relationship between bulk properties of compounds and the molecular structure of those compounds provides a connection between the macroscopic and microscopic properties of matter. Linear free energy relationship (LFER) and quantitative structure– activity/property relationships (QSAR/QSPR) have been widely adopted for estimation of behaviours of organic environmental contaminants [7– 9]. Briefly, LFER is based on the assumption that KOC (normalized to organic carbon, OC) is a function of the structure of a molecule, which in turn affects the standard thermodynamic free energy change, thus permitting the implementations of LFER, e.g. in environmental chemistry. QSAR and LFER correlations have been used both for empirically and very extensively for computationally derived descriptors. QSAR is in fact a natural extension of the LFER approach but, unfortunately, many QSAR regressions, e.g. log KOW versus a wide range of measured, theoretical or computationally derived descriptors, do not work well. Therefore, it is essential to take into account the real matrix effects, such as steric and electronic, between the sorbates and

the dissolved/solid HS sorbents, as also underlined by Lassen et al. [10]. The chemical composition, structure and reactivity of the natural HS sorbent have been, long ago, emphasized to play an important role on the extent and nature of organic contaminants [11–13]. Likewise, several authors have proposed that dissolved/solid HS sorbents may be viewed as a three-dimensional network of polymer-like chains exhibiting a relatively open, flexible structure perforated with ‘micro-voids’ [14–16]. It is also important to remember that the isolation and fractionation procedures of, e.g. aquatic humic solutes will apparently strengthen their binding affinities on organic contaminants as compared with the original situation predominating in the water system, as recently demonstrated [17] for a neutral hydrophobic organic pollutant. The maximum concentrations of different chlorophenols in natural fresh water environments are, as a rule, clearly below 1 mg l − 1 being mostly at the concentration level of ng l − 1 [18,19]. Thus, it is necessary to take into account the multiform differences between the laboratory and real natural conditions when adapting the obtained results in use. This study was undertaken to obtain more information, especially in the light of structural characterisation, about the sorption and binding mechanism of different chlorophenols with an aquatic very heterogeneous HS isolated from a brown-water lake. The selection of different chlorophenols allows the comparisons to be made between somewhat similar molecules which differ by several orders of magnitude in their octanol–water distribution and their degree of dissociation at the experimental pH. The binding capacities, equilibrium coefficients (KOC, l kg − 1 of OC) and rate constants were determined at very long contact times for four chlorophenols (2,4-DCP, 2,4,6-TCP, 2,3,4,5TeCP and PCP) at a constant HS concentration and at three different acidities using two kinds of chlorophenol compositions. The fitness of linear isotherm and non-linear Freundlich and Langmuir isotherms was also tested in describing different ion sorption models.

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2. Experimental

2.1. Reagents PCP (CAS 87-86-5) was purchased from EgaChemie/Aldrich (purity\ 99%), 2,4-DCP (CAS120-83-2) from Aldrich (purity\ 97%), 2,4,6-TCP (CAS 88-06-2) and 2,3,4,5-TeCP (CAS 4901-51-3) were donated by University of Jyva¨ skyla¨ (purity\96%). Hexane was HPLC grade. The anhydrous sodium sulfate (Merck) was pro analyse grade. High purity deionised water (Elgastat, UHQ-PS) was used for all the solutions. The humic fresh water sample was collected from Lake Savoja¨ rvi (SS, colour as cobalt– platinum units, ca. 150 mg Pt l − 1; DOC 19 mg C l − 1; pH 5.8), situated in a marsh region in the southwestern part of Finland, in September 1994. The aquatic HS (SS.[DEAE]) was isolated [20] from the lake water sample by DEAE cellulose (diethylaminoethylcellulose). It has been stated [21– 24] that the DEAE cellulose extracts practically all the organic macromolecular acids (equal to so-called humic substances) at the natural acidity of the fresh water sample. The SS.[DEAE] isolate of the present study accounted for ca. 78% of DOC containing in addition to the larger molecular size aggregates a significant amount of smaller size humic-like constituents. It has also been shown previously [20,25– 29] that the SS.[DEAE] isolate represents a more realistic humic solute fraction, especially from the structural chemistry’s point of view, of fresh water than do other chromatographic isolates which usually have been defined at strongly acidic conditions as separate humic (HA) and fulvic (FA) acid fractions whose mixture accounts for ca. 50– 55% of DOC. The SS.[DEAE] sample was after its isolation hermetically sealed and stored in the dark at 4 °C.

2.2. Sorption studies The glassware was washed with acetone, distilled water and ethanol (99%) to avoid contamination. The constant amount of the SS.[DEAE] humic sorbent (23.5 mg l − 1, on a moisture- and ash-free basis, being equivalent to 12.67 mg l − 1 of

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organic carbon, SS.[DEAE]-OC) was dissolved in every set of measurements in distilled water. The SS.[DEAE] concentration was practically analogous to the original aqueous concentration. The acidities (pH 3, 5.5 and 7) of the solutions were adjusted with sodium hydroxide or hydrochloric acid. The ionic strength was no more adjusted. The approximate ionic strength (I: 1× 10 − 4) of the solution and also the ash content (ca. 2%) of SS.[DEAE] were very low and, hence, the possibility of ion pairs being formed is almost minor. PCP, 2,3,4,5-TeCP, 2,4,6-TCP and 2,4DCP were added into the reaction mixture from their stock solutions (1 mg ml − 1 in ethanol), and thus the influence of the very slight amount of ethanol on the reaction could be ignored. In the study of binding capacities the initial concentrations for every four chlorophenols were kept constant (0.16 mg l − 1) in their individual and co-solutions. The sorption of chlorophenols onto SS.[DEAE] –OC was determined at the three acidities as a function of time (maximum contact time 190 h at 20 °C). The sorption isotherms of the four chlorophenols onto SS.[DEAE]–OC were measured in their individual solutions at pH 5.5 and at 20 °C. Sorption isotherms were generated after contact time of 170 h using seven initial concentrations for each chlorophenol (0.0005, 0.001, 0.02, 0.07, 0.14, 0.16 and 0.2 mg l − 1). All the mixtures were shaken in the dark (20 °C) on a reciprocating shaker. The samples (10 ml) taken at different intervals for the determination of the unbound chlorophenols were extracted four times with 4-ml portions of hexane. The stability of the chloro-humic adduct formed was verified in the recent PCP study [30] by adding 20 ml of hexane into the residual aquatic layer and by shaking the mixture for 2 weeks. No more free chlorophenol was found from this hexane layer confirming the stability of the formed SS.[DEAE] –chlorophenol complex. The elution of SS.[DEAE] –chlorophenol complex during the hexane extraction was verified by UV–vis spectrophotometer indicating that no detectable amount of HS was coeluted. The four hexane layers were combined, dried with anhydrous sodium sulfate and when needed concentrated with a mild nitrogen flow to the final volume. The

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accuracy of the hexane extraction was ensured by adding a known amount of chlorophenol to the purified water (practically 100% recovery). It was also verified that the presence of the four different chlorophenols in their co-solution did not noticeably influence the accuracy of the hexane extraction. Splitless injection of 1 ml of the hexane solution was carried out into a gas chromatograph (GC, Perkin Elmer Auto System) connected to an electron capture detector (ECD) for analysing the unbound chlorophenols. The capillary column was DB-5.625 (30× 0.317 mm2 i.d., film thickness 0.50 mm), which permits the direct analysis of chlorophenols without derivatization [31]. The temperature programme was 6 °C min − 1 from 100 to 280 °C. The other GC-conditions were: carrier gas helium (1.5 ml min − 1), injector and detector temperature’s 250 and 300 °C using nitrogen as make-up gas (30 ml min − 1). Identification was based on the retention times of authentic standard compounds and quantification on the calibration curves. The linear range of the EC detector was moderate between 10 and 300 pg (the change of the response was B10%). The practical quantitation limit was 51 pg.

3. Results and discussion

3.1. Binding capacity Fig. 1a–c demonstrate the binding capacities, Cs (kg of chlorophenol kg − 1 of SS.[DEAE] – OC), of the four chlorophenols (PCP, 2,3,4,5-TeCP, 2,4,6-TCP and 2,4-DCP) in their co-solutions at three different acidities onto SS.[DEAE]– OC versus the sorption time. The optimum non-linear sorption profiles for the original data points were generated by aid of the Boltzmann equation producing sigmoidal curves and the goodness of fit is given as Chi-square test. For the sake of comparison all the Cs values are expressed on the same scale at different acidities. Fig. 1a– c indicate that apparent differences occur between binding affinities, mechanisms and binding rates of different chlorophenols. These properties also vary with the acidity of the individual chlorophenol in question.

Various chlorophenols also need different times for reaching the equilibrium state. Therefore, the sorption contact time as big as 170 h was selected to estimate the maximum ‘true’ binding capacities for the individual and combined solutions of chlorophenols with different acidities. Table 1 gives the equilibrium sorption coefficients (KOC = Cs/Cb) calculated for maximum binding capacities after 170 h (20 °C) at different acidities for individual and combined solutions of different chlorophenols. The log KOC values obtained for the solutions of individual chlorophenols were systematically slightly greater (5.49 0.6%) than those obtained for their co-solutions. This indicates that some kind of ‘competition’ between chlorophenols occurs in binding onto the available sites of the aquatic HS. Table 1 demonstrates also the well-known fact [32–35] that the magnitudes of the sorption coefficients of chlorophenols are strongly pH dependent. It was recently demonstrated [30] that the magnitude of KOC depends also on the initial concentration of a chlorophenol being somewhat greater for the lower concentration ranges. The observation that sorption of chlorinated phenols by organic sorbents in natural waters is linearly dependent on chlorophenol concentrations was confirmed also long ago by Schellenberg et al. [33]. It is noteworthy, that the variation of log KOC values in the literature [8], e.g. for the PCP sorption, is very large (2.21–5.65) depending on analytical methods, algorithms applied, the nature of HS, the acidity and ionic strength of the solution, etc. The log KOC value of 3.13 at pH 5.5 for PCP in Table 1 differs somewhat from the literature values given at different environmental conditions: 3.73 [9], 3.40 [36], 2.96 [1] and 2.70 [19] (the last is an estimated ‘practical’ value for fresh water and marine environments). The log KOC value of 3.00 at pH 5.5 for 2,3,4,5-TeCP in Table 1 was significantly smaller than the average value, 3.87, given in an extensive literature review of Shiu et al. [8] fitting, however, within the range of 2.94–4.66 (n=15). The log KOC value of 2.57 at pH 5.5 for 2,4,6-TCP in Table 1 is not far from the value of 2.75 reported [1] by Seip et al. However, this experimental value is significantly smaller than

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Fig. 1. Sorption of chlorophenols onto SS.[DEAE] – OC at different acidities as a function of time (20 °C). Cs =equilibrium concentration of chlorophenol in sorbent (kg of chlorophenol kg − 1 of SS.[DEAE]– OC). Initial concentration of each chlorophenol in their co-sloutions = 0.16 mg 1 − 1 SS.[DEAE]–OC)= 12.67 mg 1 − 1.

1.555×10−4 1.195×10−4 5.022×10−5 3.533×10−5 8.112×10−5 4.093×10−5 2.499×10−5 1.572×10−5

pH 5.5 PCP 2,3,4,5-TeCP 2,4,6-TCP 2,4-DCP

pH 7 PCP 2,3,4,5-TeCP 2,4,6-TCP 2,4-DCP 1.590×10−7 1.595×10−7 1.597×10−7 1.598×10−7

1.580×10−7 1.585×10−7 1.594×10−7 1.596×10−7

1.569×10−7 1.575×10−7 1.591×10−7 1.593×10−7

5.103×102 2.567×102 1.565×102 9.839×101

9.837×102 7.542×102 3.151×102 2.214×102

1.567×103 1.243×103 4.700×102 3.411×102

2.71 2.41 2.19 1.99

2.99 2.88 2.50 2.35

3.20 3.09 2.67 2.53

1.606×10−4 5.051×10−5 2.992×10−5 2.011×10−5

2.111×10−4 1.594×10−4 5.888×10−5 4.720×10−5

3.776×10−4 3.117×10−4 1.020×10−4 7.735×10−5

1.580×10−7 1.594×10−7 1.596×10−7 1.597×10−7

1.573×10−7 1.580×10−7 1.593×10−7 1.594×10−7

1.552×10−7 1.561×10−7 1.587×10−7 1.590×10−7

Cb (kg l−1

1.017×103 3.170×102 1.874×102 1.259×102

1.342×103 1.009×103 3.697×102 2.961×102

2.433×103 1.997×103 6.428×102 4.864×102

KOC (l kg−1 of OC)

3.01 2.50 2.27 2.10

3.13 3.00 2.57 2.47

3.39 3.30 2.81 2.69

log KOC (l kg−1 of OC)

Concentration of the aquatic SS.[DEAE]–OC humic sorbent =12.67 mg l−1. Initial concentration of each chlorophenol =0.16 mg l−1. Equilibrium time = 170 h (20 °C). Sorption coefficient: KOC =Cs/Cb, Cs =equilibrium concentration of chlorophenol in sorbent and Cb = equilibrium concentration of chlorophenol in solution.

2.459×10−4 1.959×10−4 7.476×10−5 5.434×10−5

log KOC (l kg−1 of OC)

Cs (kg kg−1 of OC)

KOC (l kg−1 of OC)

Cs (kg kg−1 of OC)

Cb (kg l−1)

Individual solutions of chlorophenols

Co-solutions of chlorophenols

pH 3 PCP 2,3,4,5-TeCP 2,4,6-TCP 2,4-DCP

Chlorophenol

Table 1 Equilibrium sorption coefficients of chlorophenols for aquatic HS at different conditions

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that of 3.02 estimated [9] by Sabljic´ et al. the through linear relationship based on octanol– water partition coefficients but it is definitely within the extensive range of 1.94– 3.84 (n = 21) reviewed [8] by Shiu et al. The log KOC value of 2.47 at pH 5.5 for 2,4-DCP in Table 1 was clearly smaller than the value of 2.75 estimated [9] mathematically by Sabljic´ et al. and those of 2.84 and 2.89 presented by Seip et al. [1] and Fytianos et al. [37], respectively. It was even smaller than the average value of 2.83 calculated [8] on the extensive literature data of Shiu et al. (1.74–3.98, n =19). The above critical examination verifies that the KOC values of the present study are generally speaking reasonable. The reason for the smaller sorption constants (log KOC in Table 1) obtained in the present study compared with the literature values is probably due, in addition to environmental conditions, to the applied liquid– liquid extraction method. It is essential to keep in mind that with the liquid –liquid extraction method only the part of phenols is detected as sorbed, that was bound to the HS by strong interactions. For weaker bindings a shift of the dynamic equilibrium towards the unbound phenols takes place during the extraction. This effect contributes to the smaller sorption constants compared with those obtained by, e.g. solid phase microextraction as reported by Ohlenbusch et al. [38]. Nevertheless, the results emphasize the important role of the sorbent matrix (i.e. the content and chemical characteristics of organic carbon) for the sorption. Besides, on the contrary what has been stated [19], it is evidently not potential to try evaluate a catch-all KOC value for different chlorophenols, e.g. due to their large overlapping effects.

3.2. Sorption mechanism Table 2 demonstrates more closely the binding affinities of the four chlorophenols onto the aquatic HS in two solutions of chlorophenols with different acidities. Up to nine distinct COOH-groups with different pKa values were obtained [20] for the aquatic SS.[DEAE] sorbent. Other chemical basic characteristics of SS.[DEAE] are given, e.g. in the recent PCP study [30]. Average concentrations of protonated forms for COOH-groups and

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chlorophenols at different acidities were estimated according to De Paolis and Kukkonen [36] by the Henderson–Hasselbach equation [39]: [HA]/[A− ]= 1/(1 + 10pH − pKa). However, in reality the situation is more complicated, e.g. as the protonated form of chlorophenol sorbs, the bulk of solution distribution will change and more ionised form will be protonated. The influence of the degree of protonation of chlorinated phenols on the sorption has been thoroughly discussed by Schellenberg et al. [33]. In the present semi-quantitative study the sorption at different acidities has been postulated as a steady state sorption, and the increased concentration of the protonated chlorophenol during the sorption as compared to the original equilibrium state has been omitted. The weight-averaged pKa value of SS.[DEAE] is ca. 4.76, being practically the same as that of PCP but it is notable that it alone may produce some errors in explaining the dissociation degree of HS. The great variation of the pKa values given in the literature for a certain chlorophenol is somewhat confusing (the same is also true for log KOW and water solubility values). This peculiarity is seen by comparing the selected literature values in Table 2 with those collected [8] by Shiu et al. It is stated [33,34] that at the acidities dominating in natural waters (ca. 4–6 pH units) the overall sorption of chlorinated phenols on natural organic sorbents is in general dominated by non-dissociated species (i.e. protonated form). Furthermore, if the pH of water is not more than one unit above the pKa of the chlorinated phenol, the contribution of phenolate sorption is practically negligible. The mass balances (theoretical maximum amounts of protonated forms of chlorophenols compared with the bound amounts) estimated in Table 2 at pH 3 and especially at pH 5.5 unambiguously support this basic statement even in the case of the most acidic PCP. Table 2 supports also the well-known fact that the increasing number of chlorine substituents leads to increase in the acidity and hydrophobicity (KOW) and decrease in the solubility, and that the binding capacity is directly proportional to the hydrophobicity and inversely to the solubility. The results in Table 2 are in accord with those in literature [33] stating that sorption of PCP and TeCP is greater than that of TCP and DCP.

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Table 2 Behaviours of SS.[DEAE] and chlorophenols at different acidities meq g−1

Initial (nmol l−1)

COOH

pKa

Acid constants, 1 2 3 4 5 6 7 8 9 Total

amounts of COOH-groups and their initial concentrations per 23.5 mg l−1 of SS.[DEAE] a 2.91 0.53 12 455 5584 32 3.06 0.62 14 570 7787 53 3.28 0.55 12 925 8477 77 3.58 0.53 12 455 9861 148 3.95 0.65 15 275 13 734 419 4.39 0.53 12 455 11 967 897 4.83 0.78 18 330 18 063 3229 5.53 0.69 16 215 16 167 8387 7.31 1.41 33 135 33 133 32 630 4.76 6.29 147 815 124 773 45 872 pKa

Initial (nmol l−1)

Acid constants and initial concentrations of the four chlorophenols b PCP 4.74 [40] 602 2,3,4,5-TeCP 5.64 [41] 690 2,4,6-TCP 6.15 [33] 812 2,4-DCP 7.85 [33,42] 982 log KOW (l kg−1)

Solubility (mg Solubility l−1) (nmol l−1)

Indi6idual solutions of chlorophenols c PCP 5.12 [7,9] 14 2,3,4,5-TeCP 4.33 [7] 166 2,4,6-TCP 3.69 [7,9] 900 2,4-DCP 3.06 [7] 4500

Co-solutions of chlorophenols d PCP 2,3,4,5-TeCP 2,4,6-TCP 2,4-DCP Sum total

[43] [40] [43] [43]

52 632 711 517 4 568 528 2 760 7362

pH 3 (nmol l−1)

pH 3 (nmol l−1)

591 688 812 982

Initial (nmol l−1)

602 690 812 982

pH 5.5 (nmol l−1)

pH 5.5 (nmol l−1)

pH 7 (nmol l−1)

1 2 3 5 14 31 123 531 22 242 22 952 pH 7 (nmol l−1)

89 400 664 977

pH 3 (nmol l−1)

18 17 7 6

pH 5.5 (nmol l−1)

10 9 4 4 pH 5.5 (nmol l−1)

3 29 101 860 pH 7 (nmol l−1)

8 3 2 2

Initial (nmol l−1)

pH 3 (nmol l−1)

pH 7 (nmol l−1)

602 690 812 982

12 11 5 4

7 7 3 3

4 2 2 1

3086

32

20

9

Binding affinities of chlorophenols onto aquatic HS in their individual solutions and co-solutions at different acidities. Initial amount of each chlorophenol = 0.16 mg l−1. Amounts of different chlorophenols bound onto SS.[DEAE] were calculated after an equilibrium time of 170 h. For the sake of comparison all concentrations are expressed in same units (nmol l−1). a Calculated concentrations of different COOH-groups in protonated forms at different acidities. b Calculated theoretical amounts of chlorophenols in protonated forms at different acidities. c Literature values of octanol–water partition coefficients and water solubilities (25 °C) of chlorophenols. Initial concentrations of chlorophenols and amounts bound onto 23.5 mg l−1 of SS.[DEAE] at different acidities. d Initial concentrations of chlorophenols and amounts bound onto 23.5 mg l−1 of SS.[DEAE] at different acidities.

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According to the mass balances shown in Table 2 it is statistically evident that all the four chlorophenols can interact at pH 3 and 5.5, also in their co-solutions, via intermolecular hydrogen bonds with even the most acidic COOH-groups (pKa :2.91) of HS and the phenolic oxygens of different chlorophenols, provided that no steric hindrance is present. The formation of hydrogen bonds is most likely the dominant sorption mechanism for chlorophenols even though the sorption of the protonated chlorophenols may take place also through hydrophobic bond formation similarly to totally hydrophobic non-ionisable organic compounds [1]. It is also stated [15,16], in connection with the molecular modelling and geometry optimization of HS, that anthropogenic (e.g. chlorophenols) and some biological substances can be involved in physical trapping in the microvoids of the HS network either with intermolecular hydrogen bonds or simply through penetration. Table 2 demonstrates that at pH 3 every chlorophenol in its protonated form can theoretically form hydrogen bonds between even the most acidic COOH-group of HS. However, at pH 3 the amount of different chlorophenols bound onto SS.[DEAE] was at its best ca. 3% for PCP and no more than only ca. 0.6% for 2,4-DCP of the theoretical amounts of their protonated forms. These low proportions imply the importance of the steric influence on the sorption (i.e. the bulk of the protonated carboxylic groups of HS is not available for the sorption). The close relationship shown in Table 2 between the decreased amount of chlorophenols bound onto SS.[DEAE] at pH 5.5 and the greatly decreased concentration of the most acidic COOH-groups of HS may also support the important role of these specific COOH-groups in the sorption. The binding affinities of 2,4,6-TCP and 2,4-DCP in their individual solutions were only ca. 37% of that obtained for PCP although they were at pH 5.5 still ca. 82 and 99%, respectively, protonated affording an opportunity to intermolecular hydrogen bonds with acidic COOHgroups of HS. On the contrary, 2,3,4,5-TeCP was only ca. 58% in protonated form at pH 5.5 but its binding affinity was yet ca.87% of that obtained

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for PCP (PCP was no more than ca. 15% in protonated form at pH 5.5). This abnormality occurring between the binding affinity and the degree of protonation of the different chlorophenols apparently demonstrates that in the case of 2,4,6-TCP and 2,4-DCP the attraction of water molecules to surround chlorophenols becomes much stronger than the interactions of these chlorophenols with, e.g. certain specific acidic COOH-groups of HS, i.e. the thermodynamic gradient driving the chlorophenols out of aqueous solution is now weaker. The mass balances in Table 2 at pH 7 are totally different from those dominating at pH 3 and 5.5. Only ca. 0.5% of PCP is now, at best, protonated at the beginning of the sorption although ca. 2.3 times that amount of PCP is bound onto SS.[DEAE]. This imbalance indicates, as stated [30] previously, that at best ca. 57% of PCP must be bound at pH 7 onto SS.[DEAE] via other binding mechanisms than hydrogen or possible hydrophobic bonds, e.g. by anionic phenolate sorption, even if the freshly protonated PCP formed during the equilibrium process of the bulk solution may slightly decrease this proportion. On the other hand, all 2,3,4,5-TeCP, 2,4,6-TCP and 2,4-DCP bound onto SS.[DEAE] at pH 7 are most likely protonated. However, their interactions with water molecules will strengthen with decreasing acidity and the ability of the aqueous phase to force them to associate with HS, e.g. via hydrogen or hydrophobic bonds, becomes weaker and weaker. The association mechanism of PCP and TeCP to organic sorbents in an aqueous phase may include the solvation of ion pairs or free ions in the bulk of organic solutes and the sorption of organic ions on lipophilic surfaces with counterions in electronic double layers [44]. It is also stated [45] that neutral chlorophenols, e.g. PCP, can form conjugates with carbohydrate-like structural units. Independent of the type of binding (hydrogen or/and hydrophobic bonds or simple ‘mechanical’ penetration into the micro-voids of the HS network), the SS.[DEAE]– chlorophenol adduct is very weak and exceedingly sensitive for the acidity of the solution. It is essential to note that the quantities of chlorophenols bound onto

532

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Fig. 2. Graphic illustrations for properties of different chlorophenols’ binding onto the SS.[DEAE] humic sorbent in two kinds of mixtures with different acidities (20 °C).

the constant quantity of SS.[DEAE] at different acidities (Table 2) are very small. Even in the case of PCP the amount bound at pH 3 was only ca. 3% of the initial concentration, which in turn accounted no more than ca. 1% of the practical maximum solubility of PCP. This comparison demonstrates that the amount of DOC in the environment plays a significantly greater role than the accurate value of KOC in question.

3.3. Sorption of chlorophenols in pure solutions 6ersus their mixed systems The data given in Tables 1 and 2 are also graphically illustrated in Fig. 2 for better visualisation of the binding properties of the four different chlorophenols in the two kinds of solutions with different acidities. The decrease of the bound amounts was practically linear with increasing pH-values (i.e. decreasing acidity of the chlorophenol –HS solution) for all the chlorophenols when association occurred in the combined mixtures. On the contrary, the situation was not

so straightforward when association occurred in the solutions of single chlorophenols, especially in those of PCP, 2,3,4,5-TeCP and 2,4,6-TCP. This divergence may partly be explained by certain interactions between water molecules and hydrated forms of different chlorophenols in their co-solutions which makes the association more uniform. Fig. 2 exemplifies that the bound amount of a single chlorophenol was definitely lower in the co-solutions than in their individual solutions. On the other hand, Table 2 shows that the sum totals of bound chlorophenols in their co-solutions were considerably greater than the amount of any single chlorophenol bound from its individual solution, e.g. at pH 3 and 5.5 the sum totals were ca. twofold greater than the single amounts of PCP obtained from its pure solutions. This imbalance demonstrates, especially from the structural point of view, that the network of the humic sorbent consists of binding sites which differ from each other in their ability to bind a given chlorophenol, even though some competition will occur between various chlorophenols for certain binding sites.

J. Peura6uori et al. / Talanta 56 (2002) 523–538

3.4. Sorption kinetics and rate

533

The sorption rate constants of different chlorophenols in Table 3 were significantly greater in their individual than in their co-solutions at different acidities for the whole data set. This difference speaks also for certain competition between chlorophenols, in addition to the disturbing interactions between dissolved chlorophenols in their co-solutions, for the same binding sites of HS. This competition was most powerful for penta and tetra substituted chlorophenols being for PCP at all acidities ca. 40% faster and 2,3,4,5TeCP at pH 3 ca. 43% faster. The competition for the same binding sites was clearly less in the case of 2,4,6-TCP and 2,4-DCP at all acidities, i.e. the sorption was now only ca. 16% faster in their individual than in their co-solutions. On the other hand, a given irregularity appeared for 2,3,4,5TeCP at the acidities of pH 5.5 and 7 indicating the binding to be now most independent of prevailing conditions. At these acidities its sorption was only ca. 9% faster in its individual than in its

Table 3 shows the rate constants (k, as h − 1) estimated for the sorption of chlorophenols onto the aquatic SS.[DEAE] humic sorbent at various acidities of different mixtures. The sorption rates were calculated using a first-order kinetic equation originally derived [46] for a miscible displacement technique: d(Ct /C ) = k(C − Ct )dt; where Ct = amount of compound on the sorbent at time t, C =amount of compound on the sorbent at equilibrium, and k =sorption rate constant. This first-order kinetic function fits reasonably well (r 2 in Table 3) to the estimation of an average sorption mechanism of chlorophenols onto the aquatic HS. The sorption rate constant in Table 3 obtained, e.g. for PCP at pH 7 (0.05 h − 1) was quite close to that (0.08 h − 1) given in a literature review of Brusseau et al. [4] for PCP at pH 7.1, thus supporting the validity of the applied kinetic approach.

Table 3 Rate of sorption of chlorophenols onto SS.[DEAE] in two kinds of mixtures with different acidities No.

Chlorophenol

Mixturea

Acidity

k (h−1)

r2

log k (h−1)

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24

PCP PCP PCP PCP PCP PCP 2,3,4,5-TeCP 2,3,4,5-TeCP 2,3,4,5-TeCP 2,3,4,5-TeCP 2,3,4,5-TeCP 2,3,4,5-TeCP 2,4,6-TCP 2,4,6-TCP 2,4,6-TCP 2,4,6-TCP 2,4,6-TCP 2,4,6-TCP 2,4-DCP 2,4-DCP 2,4-DCP 2,4-DCP 2,4-DCP 2,4-DCP

* ** * ** * ** * ** * ** * ** * ** * ** * ** * ** * ** * **

PH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH pH

0.118 0.086 0.073 0.052 0.054 0.038 0.097 0.068 0.053 0.049 0.033 0.030 0.047 0.039 0.034 0.029 0.025 0.023 0.039 0.033 0.031 0.027 0.021 0.018

0.956 0.980 0.952 0.940 0.929 0.949 0.924 0.937 0.886 0.943 0.905 0.893 0.929 0.964 0.909 0.935 0.891 0.887 0.864 0.848 0.870 0.891 0.851 0.852

−0.930 −1.065 −1.136 −1.287 −1.270 −1.419 −1.015 −1.170 −1.274 −1.309 −1.479 −1.520 −1.330 −1.414 −1.469 −1.540 −1.594 −1.639 −1.404 −1.488 −1.515 −1.568 −1.684 −1.741

a

3 3 5.5 5.5 7 7 3 3 5.5 5.5 7 7 3 3 5.5 5.5 7 7 3 3 5.5 5.5 7 7

*, chlorophenol in its individual solution; **, chlorophenol in the co-solution of the four different chlorophenols.

534

J. Peura6uori et al. / Talanta 56 (2002) 523–538

co-solutions. This phenomenon of 2,3,4,5-TeCP may partly be originated in the weak inductive effect caused by the specific substitution positions of the electrophilic chlorine atoms making the aromatic ring more favourable (at the six-position) to form hydrophobic bonds, and in that way the effect of the degree of protonation will be partly compensated. Most sorption data in the literature, also in the case of hydrophobic ionisable chlorophenols, consist of a two-stage approach to equilibrium, with a rapid initial step, which accounts approximately 20–70% of total sorption, followed by a much slower one [47–50]. This tendency of the sorption (i.e. a fast kinetics followed by a slow one) can be clearly realised, e.g. at pH 3 in Fig. 1a, and the same phenomenon is also demonstrated [30] more closely in the recent PCP study. However, a simple linear equilibrium model was applied in the present study instead of a non-equilibrium approach because the very long contact time will practically allow the ‘true’ equilibrium to be achieved, as also been stated [34] elsewhere. The somewhat slow sorption kinetic values of Table 3 might be caused by the focus on the strong bindings. That makes it attractive to believe that a

weak binding of the phenols takes place with a fast kinetics and afterwards a rearrangement towards stronger bindings takes place with a slow kinetics. LFER is said [51] to exist if a linear relationship occurs when the logarithms of the rate constants of a series of compounds are plotted against the logarithms of their associated structural parameters (i.e. log k vs. log KOC). The obtained results in Tables 1 and 3 supported the log regression of the LFER relationship for the all combined data set (single and mixed solutions at all acidities, log k= 0.544 log KOC − 2.854, r 2 = 0.973, n= 24) confirming the applied simple firstorder equilibrium kinetic model connected with long contact time. This relationship indicates in general that the rate of sorption increases with the hydrophobicity/lipophilicity and the interpretations are analogous with those presented by Saarikoski et al. [42] for the sorption of phenols, anisoles and carboxylic acids.

3.5. Ion sorption models Fig. 3 shows the linear sorption isotherms of the four chlorophenols with seven different initial

Fig. 3. Isotherms for the sorption of chlorophenols by SS.[DEAE] – OC at pH 5.5 (20 °C) Cs =equlibrium concentration of chlorophenol kg 1 − 1 of SS.[DEAE]–OC. Cb = equlibrium concentration of chlorophenol in solution (unbound chlorophenol kg 1 − 1). Equlibrium time =170 h. Initial concentration of each chlorophenol in their individual sloutions: 0.0005, 0.001, 0.14, 0.16 and 0.2 mg l − 1 SS.[DEAE]–OC)=12.67 mg 1 − 1.

r2

0.964 0.984 0.943 0.968

KOC (l kg−1 of OC) 1.444×103 1.051×103 4.375×102 3.035×102

Linear isotherm

0.76 0.82 0.78 0.72

nOC

33.30 76.06 14.00 3.95

0.996 0.996 0.99 0.996

KF,OC r2 (kg(1−n) kg−1 ln)a

Freundlich isotherm

3.541×10−4 2.796×10−4 1.088×10−4 7.543×10−5

nm,OC (kg kg−1 of OC)

Langmuir isotherm

1.095×107 7.914×106 1.237×107 1.267×107

KL,OC (l kg−1 of OC)

0.953 0.891 0.982 0.937

r2

Acidity = pH 5.5. Temperature = 20 °C. Concentration of the aquatic SS.[DEAE]–OC= 12.67 mg l−1. Equilibrium time =170 h. Initial concentration range of each chlorophenol in its individual solution =0.0005–0.2 mg l−1 (seven different concentrations). a According to a dimensional analysis of the Freundlich equation the units of KF,OC vary non-linearly with nOC [61].

PCP 2,3,4,5-TeCP 2,4,6-TCP 2,4-DCP

Chlorophenols in their individual solutions

Table 4 Constants of linear, Freundlich and Langmuir isotherms for sorption of chlorophenols by SS.[DEAE]–OC

J. Peura6uori et al. / Talanta 56 (2002) 523–538 535

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concentrations in their individual solutions on the SS.[DEAE] humic sorbent after the equilibrium time of 170 h at pH 5.5. The optimum non-linear sorption profiles for the original data points were generated by aid of the Boltzmann equation. The sorption profiles in Fig. 3 indicate that the binding capacities are a function of the initial chlorophenol concentration. To reach the maximum binding capacities (Cs, max.), i.e. ‘fully saturated’ chloro-humic adduct, higher initial chlorophenol concentrations should be used, as also demonstrated [30] in the recent PCP study. Evidently, the surface of the HS contains binding sites with different binding capacities (energies), and after the strong energy sites become filled, which is reflected in the sorption profile as a decrease of the slope, sorption takes place at sites with lower binding energies (including the ‘mechanical’ penetration of the sorbate into the certain micro-voids of the HS network). Table 4 shows that the fit of the linear sorption model (Cs vs. Cb, Fig. 3) was quite good for all the four chlorophenols studied. The linear sorption coefficients (KOC =Cs/Cb) based on the best-line slopes of Fig. 3 were slightly greater than the absolute equilibrium KOC values of Table 1 (for PCP, 2,3,4,5-TeCP and 2,4-DCP ca. 5% but for 2,4,6-TCP 18%). However, the overall difference between the log KOC values was only ca. 1% which is practically insignificant. In order to decide which type of isotherm fits better than the linear model, the Freundlich (Cs = KFC nb) and Langmuir (Cs =(nmKLCb)/(1 + KLCb) equations for describing the heterogeneous sorption have been widely applied [37,52–61]. In separate cases (at low concentrations), the term KLCb in the Langmuir equation may become small as compared to 1 and the isotherm reduces to the linear form (Cs/Cb = nmKL = KOC). Accordingly, when the Freundlich exponent, n (indicating non-linearity between solution and amount adsorbed), is close to unity, the coefficient KF can be referred to as a linear distribution coefficient. However, interpretations of sorption behaviours of natural sorbents using the Freundlich isotherm may be problematic, as recently has been reminded by Carmo et al. [61]

in their mathematical review concerning unit equivalent Freundlich coefficients. Despite the exact physical meaning of the exponent, n, in the Freundlich equation is generally speaking an enigma, its adaptation offers certain so-called ‘fingerprints’ for the estimation of the sorption mechanism. In soil science the value of n is generally B 1 and it is related to the characteristics of the sorbent. It has been stated [55], based on a large selection of experimental data, that in majority of cases the value of n close to unity indicates that an L-type sorption isotherm is involved whereas n\ 1 is indicative for a co-operative sorption. In the present study the original sorption data were normalized to OC because of the common practice in the extensive literature, producing coefficients of KOC, nOC KF,OC, nm,OC and KL,OC. The different coefficients for linear, Freundlich and Langmuir isotherms have been collected in Table 4. The goodness of fit was almost equal for both the linear and Freundlich isotherms even though slight random overlapping occurred between them. On this basis, it is impossible to make a clear distinction between them. In the recent PCP study it has been demonstrated [30] that the Freundlich model was the most suitable for the low concentration range of PCP, i.e. below the initial concentration of ca. 0.001 mg l − 1 (SS.[DEAE] –OC = 12.67 mg l − 1). On the other hand, it has been reported that the linear sorption model is most suitable, e.g. for 2,4-DCP (Fytianos et al. [37]), for PCP (Warith et al. [57]) and for a carbamothioate herbicide (D’Orazio et al. [60]), and the shape [52,55] of sorption represents a C-type curve (constant partition). These apparently contradictory results emphasize the essential role of the character/ origin of natural sorbents, in addition to specific chemical properties of sorbates, in the sorption kinetics. The Freundlich nOC values (B1) in Table 4 indicated that the heterogeneity of sorption sites was the greatest for 2,4-DCP and the smallest for 2,3,4,5-TeCP. Table 4 demonstrates that the goodness of fit of the Langmuir model was also tolerable for the whole data set (except 2,3,4,5-TeCP). This outcome supports, parallel to the previous study

J. Peura6uori et al. / Talanta 56 (2002) 523–538

[30], taking into account the initial concentration range (0.0005– 0.2 mg l − 1), the idea that the Langmuir isotherm was the most suitable for the higher concentration range of PCP (ca. 0.02–1 mg l − 1) but less useful for the lower concentration ranges. The maximum theoretical binding capacity (nm,OC) indicates that ca. 58% of the binding sites of the SS.[DEAE] – OC sorbent were occupied by chlorophenols. The calculated partition coefficients (KOC =nm,OCKL,OC) were, as logarithmic units, systematically ca. 11% greater than the KOC values estimated by the linear model.

4. Conclusions (a) It is essential to take into account the structural and steric influences of natural sorbents on the sorption of anthropogenic chemicals included acidic carboxylic groups of the sorbent. (b) The structural network of aquatic humic sorbent evidently contains binding sites with different binding abilities which are able to bind best a specific chlorophenol even though some competition occurs between different chlorophenols for the same binding sites. (c) The significantly decreased (ca. 28%) affinity of an individual chlorophenol to bind onto the aquatic humic sorbent from their mixed as compared to their single solutions may partly be explained by interactions between water molecules and hydrated forms of different chlorophenols in the mixed solutions. (d) The heterogeneity of sorption sites was the greatest for 2,4-DCP and the smallest for 2,3,4,5TeCP. (e) Linear free energy relationship supports a first-order equilibrium kinetic model to be connected with the long contact time. (f) The shape of L1-type sorption isotherm fitted better in describing sorption processes than that of C-type. (g) If any specific binding mechanism or sorption isotherm is not the subject of the interest, the sorption of chlorophenols can well be described with a linear model without significant loss of the accuracy.

537

Acknowledgements The authors thank the Finnish Cultural Foundation for financial support and Professor Juha Knuutinen from the University of Jyva¨ skyla¨ for the donation of 2,4,6-TCP and 2,3,4,5-TeCP.

References [1] H.M. Seip, J. Alstad, G.E. Carlberg, K. Martinsen, R. Skaane, Sci. Total Environ. 50 (1986) 87. [2] M. Salkinoja-Salonen, M. Saxelin, J. Pere, Analysis of toxicity and biodegradability of organochlorine compounds released into the environment in bleaching effluents of kraft pulping, in: L.H. Keith (Ed.), Advanced in The Identification & Analysis of Organic Pollutants in Water, Ann Arbor Science, Michigan, 1981, pp. 1131 – 1164. [3] J. Paasivirta, H. Hakala, J. Knuutinen, T. Otollinen, J. Sa¨ rkka¨ , L. Welling, R. Paukku, R. Lammi, Chemosphere 21 (1990) 1355. [4] M.L. Brusseau, P.S.C. Rao, Chemosphere 18 (1989) 1691. [5] L.S. Lee, P.S.C. Rao, M.L. Brusseau, Environ. Sci. Technol. 25 (1991) 722. [6] M.L. Brusseau, P.S.C. Rao, Environ. Sci. Technol. 25 (1991) 1501. [7] W. Karcher, J. Devillers, Practical Applications of Quantitative Structure – Activity Relationships (QSAR) in Environmental Chemistry and Toxicology, Kluwer Academic Publishers, Dordrecht, 1990. [8] W.-Y. Shiu, K.-C. Ma, D. Varhanı´ – kova´ , D. Mackay, Chemosphere 29 (1994) 1155. [9] A. Sablji – , H. Gu¨ sten, H. Verhaar, J. Hermens, Chemosphere 31 (1995) 4489. [10] P. Lassen, L. Carlsen, Interactions between humic substances and polycyclic aromatic hydrocarbons, in: J. Drozd, S.S. Gonet, N. Senesi, J. Weber (Eds.), The Role of Humic Substances in the Ecosystems and in Environmental Protection, PTSH-Polish Society of Humic Substances, Wroclaw, Poland, 1997, pp. 703 – 708 (ISBN 83-906403-2-5). [11] N. Senesi, Sci. Total Environ. 123/124 (1992) 63. [12] N. Senesi, G. Brunetti, P. La Cava, T.M. Miano, Soil Sci. 157 (1994) 176. [13] N. Senesi, P. La Cava, T.M. Miano, J. Environ. Qual. 26 (1997) 1264. [14] R.L. Wershaw, J. Contam. Hydrol. 1 (1986) 29. [15] H.-R. Schulten, Int. J. Environ. Anal. Chem. 94 (1996) 147. [16] L.T.C. Bonten, T.C. Grotenhuis, W.H. Rulkens, Chemosphere 38 (1999) 3627. [17] J. Peuravuori, Anal. Chim. Acta 429 (2001) 77. [18] J. Peuravuori, K. Pihlaja, K. Korhonen, Analysis of Chlorinated Phenols in Water Sample —An Intercalibra-

538

[19] [20] [21] [22] [23] [24]

[25] [26] [27] [28] [29] [30] [31] [32] [33] [34] [35] [36] [37] [38] [39] [40] [41] [42]

J. Peura6uori et al. / Talanta 56 (2002) 523–538 tion Study 1-2b, Investigations Reports, University of Turku, Finland, 1995 (ISBN 951-29-0346-6). J. Muir, G. Eduljee, Sci. Total Environ. 236 (1999) 41. J. Peuravuori, K. Pihlaja, N. Va¨ lima¨ ki, Environ. Int. 23 (1997) 453. C.J. Miles, J.R. Tuschall, P.L. Brezonik, Anal. Chem. 55 (1983) 410. C. Pettersson, J. Ephraim, B. Allard, Org. Geochem. 21 (1994) 443. M.B. David, G.F. Vance, Biogeochemistry 12 (1991) 17. N. Paxe´ us, Studies on aquatic humic substances, PhD Thesis, University of Gothenburg, Sweden, 1985 (ISBN 91-7032-185-X). J. Peuravuori, K. Pihlaja, Anal. Chim. Acta 337 (1997) 133. J. Peuravuori, K. Pihlaja, Anal. Chim. Acta 363 (1998) 235. K. Hautala, J. Peuravuori, K. Pihlaja, Environ. Int. 24 (1998) 527. J. Peuravuori, N. Paaso, K. Pihlaja, Thermochim. Acta 325 (1999) 181. J. Peuravuori, N. Paaso, K. Pihlaja, Anal. Chim Acta 391 (1999) 331. J. Peuravuori, N. Paaso, K. Pihlaja, Int. J. Environ. Anal. Chem. 79 (2001) 37. M.L. Davı`, F. Gnudi, Wat. Res. 33 (1999) 3213. D.G. Crosby, K.I. Beynon, P.A. Greve, F. Korte, G.G. Stil, J.W. Vonk, Pure Appl. Chem. 53 (1981) 1051. K. Schellenberg, C. Leuenberger, R.P. Schwarzenbach, Environ. Sci. Technol. 18 (1984) 652. L.S. Lee, P.S.C. Rao, P. Nkedi-Kizza, J.J. Delfino, Environ. Sci. Technol. 24 (1990) 654. P. Lampi, K. Tolonen, T. Vartiainen, J. Tuomisto, Chemosphere 24 (1992) 1805. F. De Paolis, J. Kukkonen, Chemosphere 34 (1997) 1693. K. Fytianos, E. Voudrias, E. Kokkalis, Chemosphere 40 (2000) 3. G. Ohlenbusch, M.U. Kumke, F.H. Frimmel, Sci. Total Environ. 253 (2000) 63. D.C. Harris, Quantitative Chemical Analysis, W.H. Freeman & Co., San Francisco, CA, 1991, p. 195. K.-C. Ma, W.-Y. Shiu, D. Mackay, J. Chem. Eng. Data 38 (1993) 364. K. Ugland, E. Lundanes, T. Greibrokk, J. Chromatogr. 213 (1981) 89. J. Saarikoski, R. Lindstro¨ m, M. Tyynela¨ , Ecotoxicol. Environ. Saf. 11 (1986) 158.

[43] R. Ku¨ hne, R.-U. Ebert, F. Kleint, G. Schmidt, G. Schu¨ u¨ rmann, Chemosphere 30 (1995) 2061. [44] J.C. Westall, C. Leuenberger, R.P. Schwarzenbach, Environ. Sci. Technol. 19 (1985) 193. [45] K.R. Rao, Pentachlorophenol Chemistry, Pharmalogy and Environmental Toxicology, Plenum Press, New York, 1977. [46] D.L. Sparks, Kinetics of reactions in pure and in mixed systems, in: D.L. Sparks (Ed.), Soil Physical Chemistry, CRC Press, Boca Raton, FL, 1986, pp. 83 – 145. [47] S.W. Karickhoff, J. Hydraul. Eng. 110 (1984) 707. [48] S.W. Karickhoff, K.R. Morris, Environ. Toxicol. Chem. 4 (1985) 469. [49] G.F.R. Gilchrist, D.S. Gamble, H. Kodama, S.U. Khan, Agric. Food Chem. 41 (1993) 1748. [50] J.P. DiVincenzo, D.L. Sparks, Environ. Sci. Tech. 31 (1997) 977. [51] J. Shorter, Correlation Analysis in Organic Chemistry — An Introduction to Linear Free-Energy Relationships, Clarendon Press, Oxford, UK, 1973. [52] C.H. Giles, T.H. MacEwan, S.N. Nakhwa, D.J. Smith, Chem. Soc. (1960) 3973. [53] M. Schnitzer, S.U. Khan, Soil Organic Matter, Elsevier Science, Amsterdam, 1978, pp. 137 – 171. [54] F.J. Stevenson, Humus Chemistry —Genesis, Composition, Reactions, Wiley, New York, 1982, pp. 376 – 394. [55] G.G. Choudhry, Humic Substances —Structural, Photophysical, Photochemical and Free Radical Aspects and Interactions with Environmental Chemicals, Gordon and Breach Science Publisher, Glasgow, 1984, pp. 95 – 139. [56] W. Stumm, Chemistry of the Solid – Water Interface — Processes at the Mineral – Water and Particle – Water Interface in Natural Systems, Wiley, New York, 1992. [57] M.A. Warith, L. Fernandes, L. LaForge, Hazardous Waste Hazardous Mater. 10 (1993) 13. [58] J. Kozak, Soil organic matter as a factor influencing the fate of organic chemicals in the soil environment, in: A. Piccolo (Ed.), Humic Substances in Terrestrial Ecosystems, Elsevier Science, Amsterdam, 1996, pp. 625 – 664. [59] J.I. Drever, The Geochemistry of Natural Waters-Surface and Groundwater Environments, Prentice Hall, New Jersey, 1997, pp. 87 – 105. [60] V. D’Orazio, E. Loffredo, G. Brunetti, N. Senesi, Chemosphere 39 (1999) 183. [61] A.M. Carmo, L.S. Hundal, M.L. Thompson, Environ. Sci. Technol. 34 (2000) 4363.