The Science of the Total Environment, 123/124 (1992) 63-76 Elsevier Science Publishers B.V., Amsterdam
63
Binding mechanisms of pesticides to soil humic substances Nicola Senesi lstituto di Chimica Agraria, Universith di Bari, Via Amendola 165/.4, Bari 70126, Italy
ABSTRACT This review-paper summarizes and discusses the nature of the binding forces involved and the types of mechanisms operating, often simultaneously, in the adsorption processes of several pesticides onto soil humic substances, humic acids and fulvic acids. These include ionic, hydrogen and covalent bonding, charge-transfer or electron donor-acceptor mechanisms, Van der Waals forces, ligand exchange, and hydrophobic bonding or partitioning. Experimental evidence obtained and interpretation provided for the various adsorption processes proposed are briefly presented and commented. The review ends with some concluding remarks and recommendations for future work needed to be done. Key words: soil; humic substances; pesticides; binding mechanisms; humic acids; fulvic acids INTRODUCTION
The fate and behaviour of pesticides in the soil environment involve several different and often simultaneous phenomena including chemical, biological and photochemical degradation, transport and accumulation, volatilization and leaching, that are influenced to various extents by a number of physical, physico-chemical, biochemical, pedological and climatic factors. Several studies recently reviewed [1] have shown that, with respect to mineral components, the content and nature of organic matter (OM) in the soil play a key role in the performance of applied pesticides. The major component of soil OM is constituted by humic substances (HS), that are the most widespread and ubiquitous natural non-living organic materials occurring in all terrestrial and aquatic environments. The principal fractions of HS, humic acids (HA) and fulvic acids (FA), feature a polydispersed nature and polyelectrolytic character, surface activity properties and the presence of various chemically-reactive functional groups, free radical moieties, and hydrophilic and hydrophobic sites in their molecular structure, that qualify these substances as privileged in the interaction with organic pesticides.
64
N. SENESI
HS may interact with pesticides in several modes including adsorption, partitioning and solubilization, catalysis in hydrolysis and dealkylation, and photosensitization [1]. All these processes have evident implications in the fate and behaviour of pesticides in the soil system, affecting their degradation and detoxication, residue persistence and monitoring, mobilization and transport, bioavailability and phytotoxicity, and bioaccumulation. Adsorption phenomena represent probably the most important mode of interaction between HS and pesticides and control the concentration of the latter in the soil liquid phase. Adsorption processes may vary from complete reversibility to total irreversibility, that is, once adsorbed on HS, a pesticide may be easily desorbed, desorbed at various levels, or not at all. The extent of adsorption depends on the amount of both the HS and the pesticide and their properties, which include the size, shape, configuration, molecular structure, chemical functions, solubility, polarity, polarizability and charge distribution of interacting species, and acid or basic character of the pesticide molecule. The objective of this review is to summarize and discuss the nature of the binding forces involved and the types of mechanisms that operate, often simultaneously, in the adsorptive interaction between HS and pesticides. These include ionic, hydrogen and covalent bonding, charge-transfer and electron donor-acceptor mechanisms, Van der Waals forces, ligand exchange and hydrophobic bonding or partitioning [1]. IONIC BONDING (ION EXCHANGE)
Adsorption via ionic bonding, or cation exchange, applies only to those pesticides which are in the cationic form in solution or can accept a proton, i.e. protonate, and become cationic. It involves ionized, or easily ionizable, carboxylic and phenolic hydroxyl groups of HS. Infrared (IR) [2-5] and potentiometric titration studies [6,7] show that ion exchange is the dominant mechanism for adsorption of diquat, paraquat and chlordimeform by HS. Divalent cationic bipyridilium pesticides can react with two negatively charged sites on HS, e.g. two COO- groups or one COO- plus a phenolate ion (Fig. la). However, not all negative sites on HS seem to be positionally available to bind large organic cations, probably because of steric hindrance effects. Adsorption of phosphon and phenacridane chloride through ionic bonds onto soil OM are also reported [8]. Less basic pesticides such as s-triazines [9], amitrole [10] and dimefox [11] may become cationic through protonation depending on their basicity and the pH of the system, that also governs the degree of ionization of acidic groups on the HS. Evidence that maximum adsorption of s-triazines on
65
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
HUMIC SUBSTANCE
DIQLrAT
HLIMIC SUIISTANCE
~TRIAZINE
Rt
T
"o-
* T ' - H
I
H
RI
j,., H
a
I
H
b
Fig. 1. Ionic bond between humic substance and diquat (a) or s-triazine (b).
organic soils occurs at pH levels close to the pKa of the herbicide is indicative of ion-exchange [9]. However, the pH at the HS-surface may be of two orders of magnitude lower than that of the liquid phase [12]; thus surface-protonation of a basic molecule may occur even though the measured pH of the medium is greater than the pKa of the adsorbate. IR studies of s-triazine-HA systems show that ionic bonding can occur between a protonated secondary amino-group of the s-triazine and a carboxylate anion and, possibly, a phenolate group of the HA (Fig. lb) [13-20]. Differential thermal analysis (DTA) data confirm IR results in showing an increased thermal stability of the HA upon its interaction with various s-triazines [ 17,18]. Structural features of s-triazines, such as the nature of the substituent in the 2-position and of the alkyl groups at the 4- and 6-amino groups are shown to influence their basicity and, hence, reactivity with HS [21,22]. The higher reactivity of simazine with respect to atrazine and prometryne may be related to the smaller steric hindrance of the reactive N - H group of the former herbicide [14].
66
N. SEN~I
HYDROGEN BONDING
The presence of numerous oxygen- and hydroxyl-containing functional groups on HS renders highly probable the formation of H-bonding for pesticides containing suitable complementary groups, although a strong competition with water molecules may be expected for such sites on HS. Studies of the heat of formation of HA-atrazine complex [13,23] and several IR and DTA data [14,17,20] suggest the occurrence of one or more H-bonds, possibly involving C=O groups of HA and secondary aminogroups of the s-triazine (Fig. 2a). Acidic and anionic pesticides, such as chlorophenoxyalkanoic acids and esters, asulam and dicamba can also be adsorbed by H-bonding onto HS at pH values below their pKa in non-ionized forms through their COOH, COOR and similar groups (Fig. 2b) [15,24-27].
HUMICSUBSTANCE
s-TRIAZINE HUMICSUBSTANCE
a
HUMIC SUBSTANCE
CHLOROPHENOXYALKANOICACID
/
TC) T "%c-~c"'~°--~ ~..~-~=o---.-o / k ~C)/~--c' -/ Cl
b Fig. 2. Hydrogen bonding between humic substance and s-triazine (a) or chlorophenoxyalkanoic acid (b).
67
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
Hydrogen bonding is suggested to play an important role in the adsorption onto HS of several non-ionic polar pesticides including substituted ureas and phenylcarbamates [16,18,28,29], alachlor and cycloate [10], metolachlor [30], malathion [24] and bromacil [8]. CHARGE-TRANSFER (ELECTRON DONOR-ACCEPTOR MECHANISM)
The ascertained presence in HS of both electron-deficient structures, such as quinones, and electron-rich moieties, such as diphenols, suggests the possible formation of charge-transfer complexes, via electron donoracceptor mechanisms, with pesticides possessing, alternatively, electron donor or electron acceptor properties. The frequency shifts of the C - H out-of-plane bending vibration of paraquat, diquat and chlordimeform observed in the IR spectra of their interaction products with HS [4,32] confirm the previously postulated [2] formation
OCH=
0
OCH=
0
_
o
o
(humic quinone; electron-acccptor)
(s-triazine;
electron-donor)
(radical cation and anion; charge-transfer complex)
Huml¢ Hydroquinone (Electron Donor)
Semlquinone Radicals (Electron Donor-AeCell~OrSystem)
OH
OH
6
0
CI
CI
CI
Cl
eleclr(m Ira~d'er
0
CI
OH
Cl
014
Chloranil ( F~'¢1 t o n Acceplor)
b Fig. 3. Charge-transfer complex formation between s-triazine and humic quinone moiety (a) and chloranil and humic hydroquinone moiety (b).
68
s. SEnESl
of a charge-transfer complex between paraquat and HS (Fig. 4a). Similarly, the shift toward lower frequencies observed in the IR for the CH wagging vibration of several s-triazines upon interaction with HA [17,18,20,33,34] provides experimental evidence to the suggested formation of electron donor-acceptor complexes between methoxytriazines and soil OM [12] (Fig. 3a). These findings are confirmed by ESR studies that show an increase of the free radical concentration in the HA-s-triazine interacting system, with respect to that of unreacted HA [17,18,20,34] (Table 1). This result was explained assuming that electron-deficient, quinone-like structures in the HA remove electrons from the electron-rich donating amine and/or heterocyclic nitrogen atoms of the triazine molecule via single-electron donor-acceptor processes that involve semiquinone free radical intermediates. The enhanced conjugation possibilities offered to free electrons by the increased molecular complexity of the electron donor-acceptor system would thus result in the increased stabilization and lifetimes, that is, concentrations of free radical species. This interpretation is supported by the observed high efficiency in adsorbing s-triazines exhibited by a quinone-rich HA that also show, among various HA examined, the highest relative increase in free radical concentration, upon interaction with different s-triazines [20]. Further studies have shown that the increase of free radical concentration in various HA-striazine interaction products, that is the capacity of HA to form electron donor-acceptor systems, is inversely correlated with the content of both COOH and OH group of HA [17].
TABLE 1 Electron spin resonance (ESR) data for soil humic acid and its products of interaction with some herbicides (see Refs 18 and 25) Product of interaction of humic acid with
Free radical content (spins/g x 10 -17)
Line width factor (Gauss)
Spectroscopic splitting
Prometone Fenuron 2,4-D 2,4,5-T
23.56 33.87 5.52 5.54
8.2 6.5 6.5 6.6
2.0035 2.0033 2.0034 2.0032
Original Humic acid
8.62
6.2
2.0035
(g-values)
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
69
Charge-transfer systems such as HA-s-triazine may also originate under the effect of light, which may introduce an unpairing of electrons involved in the interaction, thus producing an increase in the ESR signal (photoinduced transfer), similar to effects observed in organic electron donoracceptor complexes of various nature [35,36]. Similar to s-triazines, substituted ureas and amitrole also possess electron donor capacity and may interact with electron-acceptor moieties in HA to form charge-transfer complexes. Evidence of this is the marked increase of free radical concentration measured in the HA-herbicide products of interaction, with respect to unreacted HA [20,10,29]. The observed positive trend between the bioactivity of the urea herbicides and the free radical concentration measured in the H A - u r e a interaction products [29] suggest that important parallels, based on similar electron donor-acceptor mechanisms and influenced by the same molecular parameters, exist between chemical reactions that these herbicides undergo with HS in soil and the biological processes involved in their herbicidal action in plants (phytotoxicity), that is, photosynthesis inhibition by interference with single-electron transfers in the Hill reaction in chloroplasts. The molecular and chemical properties of the s-triazine and substituted urea affect to some extent the efficiency in forming electron donor-acceptor systems with HA. For instance, the highest increase of free radical concentration is measured in the interaction products of various HA with prometone, among different s-triazines and with fenuron, among different substituted ureas [17,20]. These effects were ascribed to the presence of the activating, electron-donor methoxyl group on the 2-position of the ring and one isopropyl group on each amino-group of prometone and to the absence of deactivating chlorine atoms on the phenyl ring of fenuron, that would reinforce the electron donor capacity of this urea in comparison with ureas containing one or two chlorine atoms on the ring. The capacity of HS to act as electron donors in the interaction with electron acceptor pesticides and to form a charge-transfer complex has been shown for chloranyl, by direct measurement of the binding constants using ultraviolet (UV) spectroscopy (Fig. 3b) [37]. COVALENT BINDING
Formation of covalent bonds, mediated by chemical, photochemical or enzymatic catalysts and leading to stable, mostly irreversible incorporation into HS of pesticides and, more likely, of their intermediates and products of degradation (e.g., anilines and phenols), is known to occur. Acylanilides, phenylcarbamates, phenylamide, phenylureas, dinitroaniline herbicides, nitroaniline fungicides and organophosphate insecticides, such as
70
u. SENESI
parathion and methylparathion, are known to be biodegraded in soil with the release of aromatic amines, such as chloroanilines. These residues can be chemically bound to soil OM without the intervention of microbial activity, by two possible mechanisms involving carbonyl, quinone and carboxyl groups of HS and leading to hydrolyzable (probably anil, a Schiff base, and anilinoquinone) and to non-hydrolyzable (probably heterocyclic rings or ether) bound forms [38-40]. ESR studies have shown that a considerable quenching of the original free radical concentrations and a broadening of the resonance linewidths occur in a number of HA upon their interaction with water dissolved chlorophenoxyalkanoic acids and esters [25-27]. These results suggest that homolitic cross-coupling reactions, leading to incorporation through strong covalent bonds, occur between indigenous free radicals in HA and highly reactive phenoxy and/or aryloxy radicals originated as intermediates in the course of the degradation of chlorophenoxyalkanoic compounds (Fig. 4). The presence in HS of indigenous inorganic catalysts (e.g., cupric and ferric ions) and residual enzymatic activity (e.g., due to phenoloxidases), both able to mediate chemically or biologically the oxidative degradation of these herbicides, is ascertained [41]. Phenoxy-type radicals may also be generated photochemically from phenoxyalkanoic compounds in the initial oxidation step of the non-biological degradation process they undergo in aqueous solution and in presence of light and air [42,43]. The coupling reactivity of HA free radicals with chlorophenoxy-derived radicals, assumed to be proportional to the extent of lowering of the residual free radical concentration in the interaction products, has been found to be negatively correlated both with the carboxyl content and COOH to phenolic OH ratio of the HA and with the number of chlorine atoms on the phenoxy ring of the herbicide, that most likely interferes with cross-linking to the HA macromolecule [26,27].
oc,,-coo-
o(~..
"'",
+
--CO 2 CI 2.4-- D
~
(el) RADICAL INTERMEDIATE
ocH. - -
'
o
I()1
coupling ~ , ~ OH HU$tI( SEMIQL I%O~,[.
(eli
OH
|tl. M I ( I \ ( ' O R I q ) R , t I ED R I-k'~11)l. F~
Fig. 4. Formation of a radical intermediate from 2,4-D and cross-coupling reaction with a humic semiquinone radical with incorporation of the residue into the humic molecule.
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
71
The significant increase of free radical concentrations and enlargement of ESR signal linewidths measured in several HA-s-triazine interaction products [17,18,20] confirm the formation of covalent bonds, previously suggested on the basis of the increased adsorption of s-triazines onto HA observed at high temperatures [12]. These results are attributed to the increased stabilization attained by free electrons onto extended aromatic structures originated by the covalent binding of amino groups of the s-triazine to carbonyl and quinone groups of the HA. Much evidence exists of the enzyme-catalyzed incorporation into HA polymers of chloro- and alkyl-substituted aniline residues in the presence of phenoloxidases [44,45] and of chlorocatechol intermediates of the decomposition of 2,4-D and 2,4,5-T in the presence of peroxidase [45,46], whereas no evidence is available for the enzymatic binding and incorporation of striazine and phenylurea residues in HS [47]. VAN DER WAALS FORCES
Van der Waals forces consist of weak, short-range dipolar or induceddipolar attractions that operate, eventually in addition to stronger binding forces, in all adsorbent-adsorbate interactions. They assume particular importance in the adsorption of non-ionic and non-polar pesticides on suitable sites of HA molecules. Since these forces are additive, their contribution increases with the size of the interacting molecule and with its capacity to adapt to the HA surface. Although scarce experimental evidence is available, Van der Waals forces are considered to be actively involved in the adsorption onto soil HS of largesize bipyridilium cations [3], carbaryl and parathion [48], alachlor and cycloate [10], benzonitrile and DDT [49] and several thiocarbamates, carbothioates and acetanilides [8] and to represent the principal adsorption mechanism for picloram and 2,4-D [6,30]. LIGAND EXCHANGE
Adsorption by ligand exchange mechanism involves the replacement of hydration water or other weak ligands partially holding polyvalent cations associated to soil OM by suitable adsorbent molecules such as s-triazines (Fig. 5) and anionic pesticides [50], whereas it is considered unlikely in the case of linuron [51]. HYDROPHOBIC ADSORPTION AND PARTITIONING
Hydrophobic adsorption is proposed as a pH-independent mechanism for
72
N. SENESl RI
-k CO0-
~OH~
I l U m l t SUI~IIII~e - C a l l o n - W s l e r Brl~e
.~R~,/;~.~, ~Trblzln¢
- y
-coo'-" ""o., Hunll~ ~ul~lan¢¢ . I ' a l l o n - ~- 1"15~1il.. Briage
Fig. 5. Ligand exchange mechanism of adsorption of s-triazine to humic substance through a cation bridge.
retention by hydrophobic active sites of HS of non-polar pesticides that interact weakly with water and for which water molecules are not good competitors. These sites include aliphatic side-chains or lipid portions and lignin-derived moieties with high carbon content and a small number of polar groups of the HS macromolecules. Hydrophobic adsorption by soil OM and HS is suggested as an important mechanism for DDT and other organochlorine insecticides [49], leptophos, methazole, norflurazon, oxidiazinon, butralin and profluralin [15], metolachlor [30], picloram and dicamba [6], 2,4-D [24], parathion [48], phenylcarbamates [52] and substituted anilines [8], whereas it is considered a possible, additional interaction mechanism for s-triazines [53] and phenylureas [51]. Hydrophobic retention need not be an active adsorption mechanism, but can also be regarded as a partitioning between a solvent and a non-specific surface. This interaction is modeled as an equilibrium reaction, similar to the partitioning between two immiscible solvents, such as water and an organic solvent like 1-octanol. This means that HS both in the solid- and dissolvedphase are treated as a non-aqueous solvent into which the organic pesticide can partition from water [54]. For example, the apparent water solubility of a number of pesticides including DDT and lindane increased linearly with concentration of soil HA and FA [55]. CONCLUDING REMARKS AND RECOMMENDATIONS
Humic substances may interact with pesticides in different modes, of which adsorption is probably the most important one. A major question is the reversibility or irreversibility of the adsorption process, that is, whether the bound residues can be considered definitely inactivated and have become common components incorporated in the humic polymer, or they are only
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
73
momentarily inactivated in reversibly-bound forms, thus representing a possible source of contamination by a time-delayed release of toxic units. Although experimental evidence of at least partial re-mobilization of pesticide residues has been obtained as, for example, for dichloroanilinederivatives of propanil, intact methoxychlor, methylparathion, dinitroanilines and methabenzthiazuron, the question is still unresolved and needs further research. Other modes of interaction between HS and pesticides (not covered in this review) include catalytic effects of HS in hydrolysis and dealkylation reactions, photosensitization effects in various photo-degradation reactions and solubilization effects. These processes lead to the formation of intermediates or products of degradation having a mobility, toxicity and persistence which are different from those obtained in the absence of HS and that may influence the speciation of pesticides, in particular their partition, in the soil-water-organism system. The knowledge of the chemical nature, reactivity and properties of HS, the major materials interacting with pesticides in the soil, is extremely important in determining the mode and extent of the interaction. Our understanding of HS and of their multiple modes of interaction with pesticides needs further research by a more extended application of advanced techniques such as NMR, ESR, FT-IR and fluorescence spectroscopies. Adsorption of pesticides onto soil HS gives rise to a problem in the analytical, qualitative and quantitative, determination of pesticide residues in soil and water. Thus, it appears necessary to develop new procedures and methods which take into consideration these aspects and may bring a solution. Joint interdisciplinary research efforts thus appear necessary for the establishment of predictive measures to face and minimize pesticide pollution problems for soil, water, organisms and the global environment. REFERENCES 1 N. Senesi and Y. Chen, Interactions of toxic organic chemicalswith humic substances, in Z. Gerstl, Y. Chen, U. Mingelgrinand B. Yaron (Eds.), Toxic Organic Chemicalsin Porous Media, Springer-Verlag,Berlin, 1989, pp. 37-90. 2 I.G. Bums, M.H.B. Hayes and M. Stacey, Spectroscopicstudies on the mechanismsof adsorption of paraquat by humic acid and model compounds. Pestic. Sci., 4 (1973) 201-209. 3 I.G. Bums, M.H.B. Hayes and M. Stacey, Some physico-chemicalinteractions of paraquat with soil organic materials and model compounds II. Adsorption and desorption equilibria in aqueous suspensions. Weed Res., 13 (1973) 79-90. 4 S.U. Khan, Adsorption of bipyridilium herbicides by humic acids. J. Environ. Qual., 3 (1974) 202-206. 5 C. Maqueda, J.L. Perez Rodriguez, F. Martin and M.C. Hermosin, A study of the in-
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teraction between chlordimeform and humic acid from a typic chromoxevert soil. Soil Sci., 136 (1983) 75-81. S.U. Khan, Equilibrium and kinetic studies of the adsorption of 2,4-D and picloram on humic acid. Can. J. Soil Sci., 53 (1973) 429-434. D.R. Narine and R.D. Guy, Binding of diquat and paraquat to humic acid in aquatic environments. Soil Sci., 133 (1982) 356-363. J.B. Weber, Interaction of organic pesticides with particulate matter in aquatic and soil systems. Adv. Chem. Ser., 111 (1972) 55-120. J.B. Weber, S.B. Weed and T.M. Ward, Adsorption ofs-triazines by soil organic matter. Weed Sci., 17 (1969) 417-421. N. Senesi, G. Padovano, G. Loffredo and C. Testini, Interactions of amitrole, alachlor and cycloate with humic acids. Proc. 2nd Int. Conf. Environmental Contamination, Amsterdam, 1986, pp. 169-171. R.E. Grice, M.H.B. Hayes and P,R. Lundie, Adsorption of organophosphorus compounds by soil constituents and by soils. Proc. 7th Br. Insecticide and Fungicide Conf., 11 (1973) 73-81. M.H.B. Hayes, Adsorption of triazine herbicides on soil organic matter, including a short review on soil organic matter chemistry. Res. Rev., 32 (1970) 131-174. J.D. Sullivan and G.T. Felbeck, A study of the interaction of s-triazine herbicides with humic acids from three different soils. Soil Sci., 106 (1968) 42-50. R. Turski and A. Steinbrich, Studies on the possibilities of binding herbicides of triazine derivatives by humic acids. Polish J. Soil Sci., 4 (1971) 120-124. R.D. Carringer, J.B. Weber and T.J. Monaco, Adsorption-desorption of selected pesticides by organic matter and montmorillonite. J. Agric. Food Chem., 23 (1975) 569-572. N. Senesi and C. Testini, Adsorption of some nitrogenated herbicides by soil humic acids. Soil Sci., 10 (1980) 314-320. N. Senesi and C. Testini, Physico-chemical investigations of interaction mechanisms between s-triazine herbicides and soil humic acids. Geoderma, 28 (1982) 129-146. N. Senesi and C. Testini, The environmental fate of herbicides: the role of humic substances. Ecol. Bull. Stockolm, 35 (1983) 477-490. N. Kalouskova, Kinetics and mechanisms of interaction of simazine with humic acids. J. Environ. Sci. Health B., 21 (1986) 251-267. N. Senesi, C. Testini and T.M. Miano, Interaction mechanisms between humic acids of different origin and nature and electron donor herbicides: a comparative IR and ESR study. Org. Geochem., 11 (1987) 25-30. J.B. Weber, Spectrophotometrically determined ionisation constants of 1,3-alkylamino-striazines and the relationships of molecular structure and basicity. Spectrochim. Acta, 23A (1967) 458-461. J.B. Weber, Mechanism of adsorption of s-triazines by clay colloids and factors affecting plant availability. Res. Rev., 32 (1970) 93-130. G.C. Li and G.T. Felbeck, Jr., A study of the mechanism ofatrazine adsorption by humic acid from muck soil. Soil Sci., 113 (1972) 430-433. S,U. Khan, Interaction of humic acid with chlorinated phenoxyacetic and benzoic acids. Environ. Lett., 4 (1973) 141-148. N. Senesi, C. Testini and D. Metta, Binding of chlorophenoxy-alkanoic herbicides from aqueous solution by soil humic acids. Proc. Int. Conf. Environmental Contamination, London 1984, CEP Cons Ltd., Edinburgh, pp. 96-101. N. Senesi, T.M. Miano and C. Testini, Role of humic substances in the environmental chemistry of chlorinated phenoxyalkanoic acids and esters, in: L. Pawlowski, G. Alaerts
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and W.J. Lacy (Eds), Chemistry for Protection of the Environment 1985, Studies in Environmental Science 29, Elsevier, Amsterdam, 1986, pp. 183-196. N. Senesi, T.M. Miano and C. Testini, Incorporation of water dissolved chlorophenoxyalkanoic herbicides by humic acids of various origin and nature, in: G. GiovannozziSermanni and P. Nannipieri (Eds), Current Perspectives in Environmental Biogeochemistry, CNR-IPRA, Rome, 1987, pp. 295-308. P. Gaillardon, R. Calvet and J.C. Gaudry, Adsorption de quelques phenylur6es herbicides par des acides humiques. Weed Res., 20 (1980) 201-204. N. Senesi and C. Testini, Spectroscopic investigations of electron donor-acceptor processes involving organic free radicals in the adsorption of substituted urea herbicides by humic acids. Pestic. Sci., 14 (1983) 79-89. J. Kozak, Adsorption of prometryn and metholachlor by selected soil organic matter fractions. Soil Sci., 136 (1983) 94-101. S.U. Khan, Determining the role of humic substances in the fate of pesticides in the environment. J. Environ. Sci. Health B., 15 (1980) 1071-1090. S.U. Khan, Interaction of humic substances with bipyridilium herbicides. Can. J. Soil Sci., 53 (1973) 199-204. U. Mfiller-Wegener, Uber die Bindung von s-Triazinen am Huminsfiuren. Geoderma, 19 (1977) 227-235. N. Senesi, Free radicals in electron donor-acceptor reactions between a soil humic acid and photosynthesis inhibitor herbicides. Z. Pflanzen. Bodenkd., 144 (1981) 580-586. C. Lagerkrantz and M. Yhland, Photo-induced free radical reactions in the solutions of some tars and humic acid. Acta Chem. Stand., 17 (1963) 1299-1306. R. Foster, Organic Charge-Transfer Complexes, Academic Press, London, 1969, p. 470. M.C. Melcer, M.S. Zalewski, J.P. Kassett and M.A. Brisk, Nature of the binding interactions between humic substances and hydrophobic molecules. Am. Chem. Soc. Div. Environ. Chem., 27 (1) (1987) 414-416. T.S. Hsu and R. Bartha, Interaction of pesticides-derived chloroaniline residues with soil organic matter. Soil Sci., 116 (1974) 444-452. T.S. I-Isu and R. Bartha, Hydrolysable and nonhydrolysable 2,4-dichloroaniline-humus complexes and their respective rates of biodegradation. J. Agric. Food Chem., 24 (1976) 118-122. G.E. Parris, Covalent binding of aromatic amines to humates. 1. Reactions with carbonyls and quinones. Environ. Sci. Technol., 14 (1980) 1099-1106. J.M. Bollag, S.-Y. Liu and R.D. Minard, Cross-coupling of phenolic humus constituents and 2,4-dichlorophenol. Soil Sci. Soc. Am. J., 44 (1980) 52-56. D.G. Crosby, Herbicide photodecomposition, in P.C. Kearney and D.D. Kaufmann (Eds), Herbicides: Chemistry, Degradation and Mode of Action. Vol. 2, Dekker, New York, 1976, Ch. 18, pp. 835-890. R.G. Zepp, G.L. Baughman and P.F. Schlotzhauer, Comparison of photochemical behavior of various humic substances in water: sunlight-induced reactions of aquatic pollutants photosensitized by humic substances. Chemosphere, 10 (1981) 109-117. J.M. Bollag, R.D. Minard and S.-Y. Liu, Cross-linkage between anilines and phenolic humus constituents. Environ. Sci. Technol., 17 (1983) 72-80. D.F. Berry and S.A. Boyd, Reaction rates between phenolic humus constituents and anilines during cross-coupling. Soil Biol. Biocbem., 17 (1985) 631-636. D.E. Stott, J.P. Martin, D.D. Focht and K. Haider, Biodegradation, stabilization in humus and incorporation into soil biomass of 2,4-D and chlorocatechol carbons. Soil Sci. Soc. Am. J., 47 (1983) 66-70.
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47 J.M. Bollag and M.J. Loll, Incorporation of xenobiotics into soil humus. Experientia (Basel), 39 (1983) 1221-1231. 48 J.A. Lenheer and J.L, Aldrichs, A kinetic and equilibrium study of the adsorption of carbaryl and parathion upon soil organic matter surfaces. Soil Sci. Soc. Am. Proc., 35 (1971) 700-705. 49 R.H. Pierce, C.E. Olney and G.T. Felbeck, Pesticide adsorption in soils and sediments. Environ. Lett., 1 (1971) 157-172. 50 D.C. Nearpass, Adsorption ofpicloram by humic acids and humin. Soil Sci., 121 (1976) 272-277. 51 S.U. Khan and R. Mazurkevich, Adsorption of linuron on humic acid. Soil Sci., 118 (1974) 339-343. 52 G.G. Briggs, Molecular structure of herbicides and their sorption by soils. Nature, 223 (1969) 1288. 53 A. Walker and D.V. Crawford, The role of organic matter in adsorption of the triazine herbicides by soil, in: FAO/IAEA (Eds), Isotopes and Radiation in Soil Organic Matter Studies, IAEA, Vienna, 1968, pp. 91-108. 54 C.T. Chiou, P.E. Porter and D.W. Schmedding, Partition equilibria of nonionic organic compounds between soil organic matter and water. Environ. Sci. Technol., 17 (1983) 227-231. 55 C.T. Chiou, R.L. Malcolm, T.I. Brinton and D.E. Kile, Water solubility enhancement of some organic pollutants and pesticides by dissolved humic and fulvic acids. Environ. Sci. Technol., 20 (1986) 502-508.
63
Binding mechanisms of pesticides to soil humic substances Nicola Senesi lstituto di Chimica Agraria, Universith di Bari, Via Amendola 165/.4, Bari 70126, Italy
ABSTRACT This review-paper summarizes and discusses the nature of the binding forces involved and the types of mechanisms operating, often simultaneously, in the adsorption processes of several pesticides onto soil humic substances, humic acids and fulvic acids. These include ionic, hydrogen and covalent bonding, charge-transfer or electron donor-acceptor mechanisms, Van der Waals forces, ligand exchange, and hydrophobic bonding or partitioning. Experimental evidence obtained and interpretation provided for the various adsorption processes proposed are briefly presented and commented. The review ends with some concluding remarks and recommendations for future work needed to be done. Key words: soil; humic substances; pesticides; binding mechanisms; humic acids; fulvic acids INTRODUCTION
The fate and behaviour of pesticides in the soil environment involve several different and often simultaneous phenomena including chemical, biological and photochemical degradation, transport and accumulation, volatilization and leaching, that are influenced to various extents by a number of physical, physico-chemical, biochemical, pedological and climatic factors. Several studies recently reviewed [1] have shown that, with respect to mineral components, the content and nature of organic matter (OM) in the soil play a key role in the performance of applied pesticides. The major component of soil OM is constituted by humic substances (HS), that are the most widespread and ubiquitous natural non-living organic materials occurring in all terrestrial and aquatic environments. The principal fractions of HS, humic acids (HA) and fulvic acids (FA), feature a polydispersed nature and polyelectrolytic character, surface activity properties and the presence of various chemically-reactive functional groups, free radical moieties, and hydrophilic and hydrophobic sites in their molecular structure, that qualify these substances as privileged in the interaction with organic pesticides.
64
N. SENESI
HS may interact with pesticides in several modes including adsorption, partitioning and solubilization, catalysis in hydrolysis and dealkylation, and photosensitization [1]. All these processes have evident implications in the fate and behaviour of pesticides in the soil system, affecting their degradation and detoxication, residue persistence and monitoring, mobilization and transport, bioavailability and phytotoxicity, and bioaccumulation. Adsorption phenomena represent probably the most important mode of interaction between HS and pesticides and control the concentration of the latter in the soil liquid phase. Adsorption processes may vary from complete reversibility to total irreversibility, that is, once adsorbed on HS, a pesticide may be easily desorbed, desorbed at various levels, or not at all. The extent of adsorption depends on the amount of both the HS and the pesticide and their properties, which include the size, shape, configuration, molecular structure, chemical functions, solubility, polarity, polarizability and charge distribution of interacting species, and acid or basic character of the pesticide molecule. The objective of this review is to summarize and discuss the nature of the binding forces involved and the types of mechanisms that operate, often simultaneously, in the adsorptive interaction between HS and pesticides. These include ionic, hydrogen and covalent bonding, charge-transfer and electron donor-acceptor mechanisms, Van der Waals forces, ligand exchange and hydrophobic bonding or partitioning [1]. IONIC BONDING (ION EXCHANGE)
Adsorption via ionic bonding, or cation exchange, applies only to those pesticides which are in the cationic form in solution or can accept a proton, i.e. protonate, and become cationic. It involves ionized, or easily ionizable, carboxylic and phenolic hydroxyl groups of HS. Infrared (IR) [2-5] and potentiometric titration studies [6,7] show that ion exchange is the dominant mechanism for adsorption of diquat, paraquat and chlordimeform by HS. Divalent cationic bipyridilium pesticides can react with two negatively charged sites on HS, e.g. two COO- groups or one COO- plus a phenolate ion (Fig. la). However, not all negative sites on HS seem to be positionally available to bind large organic cations, probably because of steric hindrance effects. Adsorption of phosphon and phenacridane chloride through ionic bonds onto soil OM are also reported [8]. Less basic pesticides such as s-triazines [9], amitrole [10] and dimefox [11] may become cationic through protonation depending on their basicity and the pH of the system, that also governs the degree of ionization of acidic groups on the HS. Evidence that maximum adsorption of s-triazines on
65
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
HUMIC SUBSTANCE
DIQLrAT
HLIMIC SUIISTANCE
~TRIAZINE
Rt
T
"o-
* T ' - H
I
H
RI
j,., H
a
I
H
b
Fig. 1. Ionic bond between humic substance and diquat (a) or s-triazine (b).
organic soils occurs at pH levels close to the pKa of the herbicide is indicative of ion-exchange [9]. However, the pH at the HS-surface may be of two orders of magnitude lower than that of the liquid phase [12]; thus surface-protonation of a basic molecule may occur even though the measured pH of the medium is greater than the pKa of the adsorbate. IR studies of s-triazine-HA systems show that ionic bonding can occur between a protonated secondary amino-group of the s-triazine and a carboxylate anion and, possibly, a phenolate group of the HA (Fig. lb) [13-20]. Differential thermal analysis (DTA) data confirm IR results in showing an increased thermal stability of the HA upon its interaction with various s-triazines [ 17,18]. Structural features of s-triazines, such as the nature of the substituent in the 2-position and of the alkyl groups at the 4- and 6-amino groups are shown to influence their basicity and, hence, reactivity with HS [21,22]. The higher reactivity of simazine with respect to atrazine and prometryne may be related to the smaller steric hindrance of the reactive N - H group of the former herbicide [14].
66
N. SEN~I
HYDROGEN BONDING
The presence of numerous oxygen- and hydroxyl-containing functional groups on HS renders highly probable the formation of H-bonding for pesticides containing suitable complementary groups, although a strong competition with water molecules may be expected for such sites on HS. Studies of the heat of formation of HA-atrazine complex [13,23] and several IR and DTA data [14,17,20] suggest the occurrence of one or more H-bonds, possibly involving C=O groups of HA and secondary aminogroups of the s-triazine (Fig. 2a). Acidic and anionic pesticides, such as chlorophenoxyalkanoic acids and esters, asulam and dicamba can also be adsorbed by H-bonding onto HS at pH values below their pKa in non-ionized forms through their COOH, COOR and similar groups (Fig. 2b) [15,24-27].
HUMICSUBSTANCE
s-TRIAZINE HUMICSUBSTANCE
a
HUMIC SUBSTANCE
CHLOROPHENOXYALKANOICACID
/
TC) T "%c-~c"'~°--~ ~..~-~=o---.-o / k ~C)/~--c' -/ Cl
b Fig. 2. Hydrogen bonding between humic substance and s-triazine (a) or chlorophenoxyalkanoic acid (b).
67
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
Hydrogen bonding is suggested to play an important role in the adsorption onto HS of several non-ionic polar pesticides including substituted ureas and phenylcarbamates [16,18,28,29], alachlor and cycloate [10], metolachlor [30], malathion [24] and bromacil [8]. CHARGE-TRANSFER (ELECTRON DONOR-ACCEPTOR MECHANISM)
The ascertained presence in HS of both electron-deficient structures, such as quinones, and electron-rich moieties, such as diphenols, suggests the possible formation of charge-transfer complexes, via electron donoracceptor mechanisms, with pesticides possessing, alternatively, electron donor or electron acceptor properties. The frequency shifts of the C - H out-of-plane bending vibration of paraquat, diquat and chlordimeform observed in the IR spectra of their interaction products with HS [4,32] confirm the previously postulated [2] formation
OCH=
0
OCH=
0
_
o
o
(humic quinone; electron-acccptor)
(s-triazine;
electron-donor)
(radical cation and anion; charge-transfer complex)
Huml¢ Hydroquinone (Electron Donor)
Semlquinone Radicals (Electron Donor-AeCell~OrSystem)
OH
OH
6
0
CI
CI
CI
Cl
eleclr(m Ira~d'er
0
CI
OH
Cl
014
Chloranil ( F~'¢1 t o n Acceplor)
b Fig. 3. Charge-transfer complex formation between s-triazine and humic quinone moiety (a) and chloranil and humic hydroquinone moiety (b).
68
s. SEnESl
of a charge-transfer complex between paraquat and HS (Fig. 4a). Similarly, the shift toward lower frequencies observed in the IR for the CH wagging vibration of several s-triazines upon interaction with HA [17,18,20,33,34] provides experimental evidence to the suggested formation of electron donor-acceptor complexes between methoxytriazines and soil OM [12] (Fig. 3a). These findings are confirmed by ESR studies that show an increase of the free radical concentration in the HA-s-triazine interacting system, with respect to that of unreacted HA [17,18,20,34] (Table 1). This result was explained assuming that electron-deficient, quinone-like structures in the HA remove electrons from the electron-rich donating amine and/or heterocyclic nitrogen atoms of the triazine molecule via single-electron donor-acceptor processes that involve semiquinone free radical intermediates. The enhanced conjugation possibilities offered to free electrons by the increased molecular complexity of the electron donor-acceptor system would thus result in the increased stabilization and lifetimes, that is, concentrations of free radical species. This interpretation is supported by the observed high efficiency in adsorbing s-triazines exhibited by a quinone-rich HA that also show, among various HA examined, the highest relative increase in free radical concentration, upon interaction with different s-triazines [20]. Further studies have shown that the increase of free radical concentration in various HA-striazine interaction products, that is the capacity of HA to form electron donor-acceptor systems, is inversely correlated with the content of both COOH and OH group of HA [17].
TABLE 1 Electron spin resonance (ESR) data for soil humic acid and its products of interaction with some herbicides (see Refs 18 and 25) Product of interaction of humic acid with
Free radical content (spins/g x 10 -17)
Line width factor (Gauss)
Spectroscopic splitting
Prometone Fenuron 2,4-D 2,4,5-T
23.56 33.87 5.52 5.54
8.2 6.5 6.5 6.6
2.0035 2.0033 2.0034 2.0032
Original Humic acid
8.62
6.2
2.0035
(g-values)
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
69
Charge-transfer systems such as HA-s-triazine may also originate under the effect of light, which may introduce an unpairing of electrons involved in the interaction, thus producing an increase in the ESR signal (photoinduced transfer), similar to effects observed in organic electron donoracceptor complexes of various nature [35,36]. Similar to s-triazines, substituted ureas and amitrole also possess electron donor capacity and may interact with electron-acceptor moieties in HA to form charge-transfer complexes. Evidence of this is the marked increase of free radical concentration measured in the HA-herbicide products of interaction, with respect to unreacted HA [20,10,29]. The observed positive trend between the bioactivity of the urea herbicides and the free radical concentration measured in the H A - u r e a interaction products [29] suggest that important parallels, based on similar electron donor-acceptor mechanisms and influenced by the same molecular parameters, exist between chemical reactions that these herbicides undergo with HS in soil and the biological processes involved in their herbicidal action in plants (phytotoxicity), that is, photosynthesis inhibition by interference with single-electron transfers in the Hill reaction in chloroplasts. The molecular and chemical properties of the s-triazine and substituted urea affect to some extent the efficiency in forming electron donor-acceptor systems with HA. For instance, the highest increase of free radical concentration is measured in the interaction products of various HA with prometone, among different s-triazines and with fenuron, among different substituted ureas [17,20]. These effects were ascribed to the presence of the activating, electron-donor methoxyl group on the 2-position of the ring and one isopropyl group on each amino-group of prometone and to the absence of deactivating chlorine atoms on the phenyl ring of fenuron, that would reinforce the electron donor capacity of this urea in comparison with ureas containing one or two chlorine atoms on the ring. The capacity of HS to act as electron donors in the interaction with electron acceptor pesticides and to form a charge-transfer complex has been shown for chloranyl, by direct measurement of the binding constants using ultraviolet (UV) spectroscopy (Fig. 3b) [37]. COVALENT BINDING
Formation of covalent bonds, mediated by chemical, photochemical or enzymatic catalysts and leading to stable, mostly irreversible incorporation into HS of pesticides and, more likely, of their intermediates and products of degradation (e.g., anilines and phenols), is known to occur. Acylanilides, phenylcarbamates, phenylamide, phenylureas, dinitroaniline herbicides, nitroaniline fungicides and organophosphate insecticides, such as
70
u. SENESI
parathion and methylparathion, are known to be biodegraded in soil with the release of aromatic amines, such as chloroanilines. These residues can be chemically bound to soil OM without the intervention of microbial activity, by two possible mechanisms involving carbonyl, quinone and carboxyl groups of HS and leading to hydrolyzable (probably anil, a Schiff base, and anilinoquinone) and to non-hydrolyzable (probably heterocyclic rings or ether) bound forms [38-40]. ESR studies have shown that a considerable quenching of the original free radical concentrations and a broadening of the resonance linewidths occur in a number of HA upon their interaction with water dissolved chlorophenoxyalkanoic acids and esters [25-27]. These results suggest that homolitic cross-coupling reactions, leading to incorporation through strong covalent bonds, occur between indigenous free radicals in HA and highly reactive phenoxy and/or aryloxy radicals originated as intermediates in the course of the degradation of chlorophenoxyalkanoic compounds (Fig. 4). The presence in HS of indigenous inorganic catalysts (e.g., cupric and ferric ions) and residual enzymatic activity (e.g., due to phenoloxidases), both able to mediate chemically or biologically the oxidative degradation of these herbicides, is ascertained [41]. Phenoxy-type radicals may also be generated photochemically from phenoxyalkanoic compounds in the initial oxidation step of the non-biological degradation process they undergo in aqueous solution and in presence of light and air [42,43]. The coupling reactivity of HA free radicals with chlorophenoxy-derived radicals, assumed to be proportional to the extent of lowering of the residual free radical concentration in the interaction products, has been found to be negatively correlated both with the carboxyl content and COOH to phenolic OH ratio of the HA and with the number of chlorine atoms on the phenoxy ring of the herbicide, that most likely interferes with cross-linking to the HA macromolecule [26,27].
oc,,-coo-
o(~..
"'",
+
--CO 2 CI 2.4-- D
~
(el) RADICAL INTERMEDIATE
ocH. - -
'
o
I()1
coupling ~ , ~ OH HU$tI( SEMIQL I%O~,[.
(eli
OH
|tl. M I ( I \ ( ' O R I q ) R , t I ED R I-k'~11)l. F~
Fig. 4. Formation of a radical intermediate from 2,4-D and cross-coupling reaction with a humic semiquinone radical with incorporation of the residue into the humic molecule.
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
71
The significant increase of free radical concentrations and enlargement of ESR signal linewidths measured in several HA-s-triazine interaction products [17,18,20] confirm the formation of covalent bonds, previously suggested on the basis of the increased adsorption of s-triazines onto HA observed at high temperatures [12]. These results are attributed to the increased stabilization attained by free electrons onto extended aromatic structures originated by the covalent binding of amino groups of the s-triazine to carbonyl and quinone groups of the HA. Much evidence exists of the enzyme-catalyzed incorporation into HA polymers of chloro- and alkyl-substituted aniline residues in the presence of phenoloxidases [44,45] and of chlorocatechol intermediates of the decomposition of 2,4-D and 2,4,5-T in the presence of peroxidase [45,46], whereas no evidence is available for the enzymatic binding and incorporation of striazine and phenylurea residues in HS [47]. VAN DER WAALS FORCES
Van der Waals forces consist of weak, short-range dipolar or induceddipolar attractions that operate, eventually in addition to stronger binding forces, in all adsorbent-adsorbate interactions. They assume particular importance in the adsorption of non-ionic and non-polar pesticides on suitable sites of HA molecules. Since these forces are additive, their contribution increases with the size of the interacting molecule and with its capacity to adapt to the HA surface. Although scarce experimental evidence is available, Van der Waals forces are considered to be actively involved in the adsorption onto soil HS of largesize bipyridilium cations [3], carbaryl and parathion [48], alachlor and cycloate [10], benzonitrile and DDT [49] and several thiocarbamates, carbothioates and acetanilides [8] and to represent the principal adsorption mechanism for picloram and 2,4-D [6,30]. LIGAND EXCHANGE
Adsorption by ligand exchange mechanism involves the replacement of hydration water or other weak ligands partially holding polyvalent cations associated to soil OM by suitable adsorbent molecules such as s-triazines (Fig. 5) and anionic pesticides [50], whereas it is considered unlikely in the case of linuron [51]. HYDROPHOBIC ADSORPTION AND PARTITIONING
Hydrophobic adsorption is proposed as a pH-independent mechanism for
72
N. SENESl RI
-k CO0-
~OH~
I l U m l t SUI~IIII~e - C a l l o n - W s l e r Brl~e
.~R~,/;~.~, ~Trblzln¢
- y
-coo'-" ""o., Hunll~ ~ul~lan¢¢ . I ' a l l o n - ~- 1"15~1il.. Briage
Fig. 5. Ligand exchange mechanism of adsorption of s-triazine to humic substance through a cation bridge.
retention by hydrophobic active sites of HS of non-polar pesticides that interact weakly with water and for which water molecules are not good competitors. These sites include aliphatic side-chains or lipid portions and lignin-derived moieties with high carbon content and a small number of polar groups of the HS macromolecules. Hydrophobic adsorption by soil OM and HS is suggested as an important mechanism for DDT and other organochlorine insecticides [49], leptophos, methazole, norflurazon, oxidiazinon, butralin and profluralin [15], metolachlor [30], picloram and dicamba [6], 2,4-D [24], parathion [48], phenylcarbamates [52] and substituted anilines [8], whereas it is considered a possible, additional interaction mechanism for s-triazines [53] and phenylureas [51]. Hydrophobic retention need not be an active adsorption mechanism, but can also be regarded as a partitioning between a solvent and a non-specific surface. This interaction is modeled as an equilibrium reaction, similar to the partitioning between two immiscible solvents, such as water and an organic solvent like 1-octanol. This means that HS both in the solid- and dissolvedphase are treated as a non-aqueous solvent into which the organic pesticide can partition from water [54]. For example, the apparent water solubility of a number of pesticides including DDT and lindane increased linearly with concentration of soil HA and FA [55]. CONCLUDING REMARKS AND RECOMMENDATIONS
Humic substances may interact with pesticides in different modes, of which adsorption is probably the most important one. A major question is the reversibility or irreversibility of the adsorption process, that is, whether the bound residues can be considered definitely inactivated and have become common components incorporated in the humic polymer, or they are only
BINDING OF PESTICIDES TO SOIL HUMIC SUBSTANCES
73
momentarily inactivated in reversibly-bound forms, thus representing a possible source of contamination by a time-delayed release of toxic units. Although experimental evidence of at least partial re-mobilization of pesticide residues has been obtained as, for example, for dichloroanilinederivatives of propanil, intact methoxychlor, methylparathion, dinitroanilines and methabenzthiazuron, the question is still unresolved and needs further research. Other modes of interaction between HS and pesticides (not covered in this review) include catalytic effects of HS in hydrolysis and dealkylation reactions, photosensitization effects in various photo-degradation reactions and solubilization effects. These processes lead to the formation of intermediates or products of degradation having a mobility, toxicity and persistence which are different from those obtained in the absence of HS and that may influence the speciation of pesticides, in particular their partition, in the soil-water-organism system. The knowledge of the chemical nature, reactivity and properties of HS, the major materials interacting with pesticides in the soil, is extremely important in determining the mode and extent of the interaction. Our understanding of HS and of their multiple modes of interaction with pesticides needs further research by a more extended application of advanced techniques such as NMR, ESR, FT-IR and fluorescence spectroscopies. Adsorption of pesticides onto soil HS gives rise to a problem in the analytical, qualitative and quantitative, determination of pesticide residues in soil and water. Thus, it appears necessary to develop new procedures and methods which take into consideration these aspects and may bring a solution. Joint interdisciplinary research efforts thus appear necessary for the establishment of predictive measures to face and minimize pesticide pollution problems for soil, water, organisms and the global environment. REFERENCES 1 N. Senesi and Y. Chen, Interactions of toxic organic chemicalswith humic substances, in Z. Gerstl, Y. Chen, U. Mingelgrinand B. Yaron (Eds.), Toxic Organic Chemicalsin Porous Media, Springer-Verlag,Berlin, 1989, pp. 37-90. 2 I.G. Bums, M.H.B. Hayes and M. Stacey, Spectroscopicstudies on the mechanismsof adsorption of paraquat by humic acid and model compounds. Pestic. Sci., 4 (1973) 201-209. 3 I.G. Bums, M.H.B. Hayes and M. Stacey, Some physico-chemicalinteractions of paraquat with soil organic materials and model compounds II. Adsorption and desorption equilibria in aqueous suspensions. Weed Res., 13 (1973) 79-90. 4 S.U. Khan, Adsorption of bipyridilium herbicides by humic acids. J. Environ. Qual., 3 (1974) 202-206. 5 C. Maqueda, J.L. Perez Rodriguez, F. Martin and M.C. Hermosin, A study of the in-
74
6 7 8 9 10
11
12 13 14 15
16 17 18 19 20
21
22 23 24 25
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N. SENESI
teraction between chlordimeform and humic acid from a typic chromoxevert soil. Soil Sci., 136 (1983) 75-81. S.U. Khan, Equilibrium and kinetic studies of the adsorption of 2,4-D and picloram on humic acid. Can. J. Soil Sci., 53 (1973) 429-434. D.R. Narine and R.D. Guy, Binding of diquat and paraquat to humic acid in aquatic environments. Soil Sci., 133 (1982) 356-363. J.B. Weber, Interaction of organic pesticides with particulate matter in aquatic and soil systems. Adv. Chem. Ser., 111 (1972) 55-120. J.B. Weber, S.B. Weed and T.M. Ward, Adsorption ofs-triazines by soil organic matter. Weed Sci., 17 (1969) 417-421. N. Senesi, G. Padovano, G. Loffredo and C. Testini, Interactions of amitrole, alachlor and cycloate with humic acids. Proc. 2nd Int. Conf. Environmental Contamination, Amsterdam, 1986, pp. 169-171. R.E. Grice, M.H.B. Hayes and P,R. Lundie, Adsorption of organophosphorus compounds by soil constituents and by soils. Proc. 7th Br. Insecticide and Fungicide Conf., 11 (1973) 73-81. M.H.B. Hayes, Adsorption of triazine herbicides on soil organic matter, including a short review on soil organic matter chemistry. Res. Rev., 32 (1970) 131-174. J.D. Sullivan and G.T. Felbeck, A study of the interaction of s-triazine herbicides with humic acids from three different soils. Soil Sci., 106 (1968) 42-50. R. Turski and A. Steinbrich, Studies on the possibilities of binding herbicides of triazine derivatives by humic acids. Polish J. Soil Sci., 4 (1971) 120-124. R.D. Carringer, J.B. Weber and T.J. Monaco, Adsorption-desorption of selected pesticides by organic matter and montmorillonite. J. Agric. Food Chem., 23 (1975) 569-572. N. Senesi and C. Testini, Adsorption of some nitrogenated herbicides by soil humic acids. Soil Sci., 10 (1980) 314-320. N. Senesi and C. Testini, Physico-chemical investigations of interaction mechanisms between s-triazine herbicides and soil humic acids. Geoderma, 28 (1982) 129-146. N. Senesi and C. Testini, The environmental fate of herbicides: the role of humic substances. Ecol. Bull. Stockolm, 35 (1983) 477-490. N. Kalouskova, Kinetics and mechanisms of interaction of simazine with humic acids. J. Environ. Sci. Health B., 21 (1986) 251-267. N. Senesi, C. Testini and T.M. Miano, Interaction mechanisms between humic acids of different origin and nature and electron donor herbicides: a comparative IR and ESR study. Org. Geochem., 11 (1987) 25-30. J.B. Weber, Spectrophotometrically determined ionisation constants of 1,3-alkylamino-striazines and the relationships of molecular structure and basicity. Spectrochim. Acta, 23A (1967) 458-461. J.B. Weber, Mechanism of adsorption of s-triazines by clay colloids and factors affecting plant availability. Res. Rev., 32 (1970) 93-130. G.C. Li and G.T. Felbeck, Jr., A study of the mechanism ofatrazine adsorption by humic acid from muck soil. Soil Sci., 113 (1972) 430-433. S,U. Khan, Interaction of humic acid with chlorinated phenoxyacetic and benzoic acids. Environ. Lett., 4 (1973) 141-148. N. Senesi, C. Testini and D. Metta, Binding of chlorophenoxy-alkanoic herbicides from aqueous solution by soil humic acids. Proc. Int. Conf. Environmental Contamination, London 1984, CEP Cons Ltd., Edinburgh, pp. 96-101. N. Senesi, T.M. Miano and C. Testini, Role of humic substances in the environmental chemistry of chlorinated phenoxyalkanoic acids and esters, in: L. Pawlowski, G. Alaerts
BINDING OF PESTICIDESTO SOIL HUMIC SUBSTANCES
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28 29
30 31 32 33 34 35 36 37
38 39
40 41 42
43
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