Importance of heavy metal-organic matter interactions in natural waters

Geochimica et Cosmochimica Acta. 1977. Vol. 41. pp. 325 to 334. Prrgamon Press Prmted m Great Bntam

Importance of heavy metal-organic matter interactions in natural waters* J. H. REUTER

School of Geophysical Sciences, Georgia Institute of Technology, Atlanta, Georgia 30332, U.S.A. and E. M. PERDUE Department of Chemistry, Portland State University, Portland, Oregon (Received

26 February

1976; accepted

in ret&d

fortn

30 August

97207. U.S.A. 1976)

Abstract-The role of organic ligands in metal complexing in natural waters has received little attention because of uncertainties regarding both the abundance and nature of dissolved organic carbon compounds. Recent data show that the bulk of dissolved organic matter in natural waters consists of highly oxidized and chemically and biologically stable polymeric compounds closely resembling soil humic substances. Average molar concentrations of these aquatic humics in major U.S. rivers range from 5 x 1O-6 to 3 x 10m5 moles l- ‘, Fractional elution of soil organic matter by meteoric waters may be considered to be the main process contributing to the presence of humic matter in rivers. The aquatic humic polymers participate in complex formation through ionizable functional groups with a range of differential acidities. The stabilities of metal-humic complexes in natural waters are higher than those of the corresponding inorganic metal complexes. Quantitative evaluation of the metal-organic interactions can be approached by applying variable equilibrium functions which take into account the differential physico-chemical characteristics of the active complexing sites on the polymer molecule. Assuming an average humic concentration of 10 mg I-‘. complexation of trace

metals can be significant even in the presence of excess concentrations

of major cations.

compounds are probably quantitatively insignificant in natural aqueous systems. In order to be able to evaluate the importance of metal-organic interactions in natural waters, one must therefore first investigate the following parameters: (1) the abundance of organic matter in natural waters, (2) the molecular nature of dissolved organic compounds; in view of the complexity of composition, the analysis of compound classes may serve as a first approximation, (3) the nature of the metal-organic interactions of the predominant compound class with heavy metals. It is the aim of this paper to provide an outline of the available information on the above parameters in fresh water environments. In estuarine and marine waters the relative abundance of inorganic ligands becomes overwhelming while the concentration of dissolved organic matter decreases rapidly as a consequence of dilution and flocculation at the freshwaterseawater interface (SHOLKOVITZ,1976).

INTRODUCTION

these

ENVIRONMENTALconcern for water resources has in recent years led to intensified research on the chemistry of natural waters. The application of quantitative models for the prediction of the behavior of heavy metals has received special attention. Most thermodynamic models extensively treat the role of inorganic ligands and exclude organic ligands. There are several reasons for this exclusion: (1) it is generally assumed that organic matter in natural waters plays a quantitatively negligible role. For example, STUMM and MORGAN (1970) state that concentrations of organic matter range from 0.1 to 10 mg I-‘, emphasizing that the upper value is restricted to polluted systems. (2) From the assumed low concentration of organic matter it follows that the ratio of the major cationic constitutents to organic ligands is very high. Thus, it is argued, the competition for complexing sites is overwhelmingly dominated by these major cations, so that complexation of trace metals becomes insignificant. (3) It is recognized that the introduction of organic ligands into thermodynamic models is not yet realistic, since the chemical nature of the naturally occurring organic molecules is very complex and still rather poorly described. On the other hand, attempts have been made to incorporate well characterized, fairly simple organic compounds into thermodynamic models (e.g. LERMAN and CHILDS, 1973), even though

COMPOSITION OF CARBON

TOTAL (TOC)

ORGANIC

Most organic matter analyses of river and lake waters are reported as total organic carbon (TOC) concentrations (see e.g. U.S. Geol. Suru. Water Resources Data for Oregon, 1974). Since an unknown fraction of the TOC occurs in particulate form, it is sometimes difficult to obtain a representative aliquot from an unfiltered water sample. Therefore, since TOC data are of somewhat limited value (Malcolm, personal communication). many researchers have

* Paper presented at the Geochemical Society Symposium ‘Sorption Versus Solubility Controls on Heavy Metals in Aquatic Environments’, Salt Lake City, October 1975. 325

326

J. H. REUTERand E. M. PERDUE

attempted to divide TOC into dissolved (DOC) and particulate (POC) fractions. In the absence of any evidence for a natural discontinuity in particle sizes, a particle size of 0.45 pm has been arbitrarily chosen as the boundary between DOC and POC. According to WETZEL and RICH (1973), the DOC/POC ratio varies from 6: 1 to 10: 1 in lakes and rivers, with the lower ratio occurring mostly in highly productive lake waters. MALCOLMand DURUM (1976) have reported DOC/POC ratios ranging from approximately 1: 1 to 18 : 1 for a variety of rivers. Although the DOC/POC approach is more satisfactory from an analytical point of view, it is difficult to avoid the erroneous concept of a discontinuity of physical and chemical properties at the DOC/POC boundary. LAMAR(1968) has shown that filtration of river water through a series of filters of progressively smaller pore size (4-0.01 pm) results in continuous removal of organic matter and iron, indicating that the relative abundance of DOC and POC (as defined) are of little chemical significance insofar as metal complexation is concerned. The compositions of DOC and POC in rivers are determined mostly by the input of allochthonous organic matter (WETZEL, 1975) which includes biopolymers derived from the decay of plant matter and geopolymers (humic substances) generated by the random polymerization of biomonomers. The biopolymers are predominantly polypeptides and polysaccharides. Since both compound classes are easily degradable by the most common hydrolytic enzyme systems, their concentrations remain low. SEMENOV et al. (1967), in a survey of organic substances in some river waters of the Soviet Union, found proteinaceous matter to constitute less than 10% of dissolved organic matter. The hydrolytic breakdown products, i.e. free amino acids, are always found at extremely low concentrations, since they are rapidly assimilated. They constitute less than 1% of the dissolved organic carbon. Dissolved polysaccharides often occur in close association with dissolved colored geopolymers (humic substances) and can make up close to 10% of dissolved organic matter (LEENHEER and MALCOLM, 1973; SEMENOV et al., 1967). Individual carbohydrate monomers, resulting from hydrolysis of polysaccharides, are almost immediately metabolized. Therefore, the instantaneous concentrations of simple sugars remain very low. Simple organic acids are subject to the same conditions as carbohydrates, and their abundance is therefore very low. Thus, on the average, about 6G80°? of DOC and POC consist of humic substances. At DOC concentrations greater than 20 mg l-l, humic substances become conspicuous to the naked eye by imparting a yellow color to natural waters. The chemistry of these color producing substances has been investigated by Christman and coworkers (CHRISTMAN,1970; CHRISTMANand MINEAR, 1971). Purified extracts of humic substances from lake and river waters were subjected to oxidative and reductive

degradations. The breakdown products consisted of mixtures of polyphenols and phenolic acids. According to CHRISTMAN (1970), the ‘color molecules’ consist of a core structure formed by the aromatic acids, to which the phenols are bound by ester and ether linkages. In comparison, oxidative and reductive degradation studies on soil humic substances have led to essentially the same breakdown products (summarized in SCHNITZERand KHAN, 1972, pp. 144 ff. and pp. 167 ff). Comparisons of the structural interpretations for the soil humic substances and the aquatic ‘color molecules’ (CHRISTMANand MINEAR, 1971) show that both are chemically and genetically closely related. In both cases, the precursor molecules appear to be derived from lignin, plant phenols (flavones, stilbenes, tannins), and phenolic metabolites of microorganisms. In addition, CHRISTMAN(1970) reports the presence of significant concentrations of amino acids in the breakdown products of aquatic color molecules. BECK et al. (1974) found that amino acids released by acid hydrolysis from river water humic substances constitute 4.4% of the organic matter. Amino acids released from soil humic substances by acid hydrolysis can constitute up to 10% of the humic material (FELBECK, 1971). The relative molar distribution of amino acids in the acid hydrolyzates of river water humic matter (BECK et al., 1974) closely resembles that found for soil fulvic acids (KHAN and SOWDEN,1971, 1972), in that glycine is the most abundant amino acid, followed by aspartic acid and alanine. These three amino acids constitute 44 and 42% of the total amino acids in river water and soil humic matter, respectively. BECK et al. (1974) and, more recently, REUTERet al. (in preparation) have analyzed dissolved river water humic matter isolated by freeze-drying for the following parameters: oxygen containing functional groups, elemental composition, i.r. absorption, and number average molecular weight. A comparison (Fig. 1) of oxygen containing functional groups (total acidity, carboxyl, and phenolic hydroxyl functions) shows that the values for unfractionated river water organic matter fall well within the range reported for soil fulvic acids. In addition, fractionation by gel permeation chromatography (GPC) yields a high molecular weight fraction which closely resembles soil humic acid (Fig. 1). A comparison of elemental composition (C, H, N, 0) shows the same similarities: average carbon content for the unfractionated aquatic humics is only slightly higher than for fulvic acids (49.1 and 47.3%, respectively) and oxygen is slightly lower (45.1 and 46.7x, respectively). The high molecular weight GPC fraction of aquatic humics has carbon and oxygen contents ranging into values characteristic for humic acids (53.3 and 41.1%, respectively). Infra-red spectra of river water humic substances show characteristic absorption bands which are the same as those of soil humic substances. According to the spectral classification suggested by STEVENSON

327

Importance of heavy metal-organic matter I @Total Acldlty

banks where water velocity is low. These aggregates are partially retained by b-45 pm filters, indicating a possible transition from DOC to POC. The origin of the high molecular weight fraction can at this time only be a matter of speculation: it is either originally coextracted with the low molecular weight materials, or it is derived from the low molecular weight humic fraction by continuing polymerization in solution (BECKet al., 1974), or derived by both processes. Solubilization of high molecular weight, humic acid type material from soil organic matter by meteoric waters (see discussion below) may be aided by the presence of rain- or weathering-derived sodium salts. Extraction of peat soils with distilled water in our laboratory has yielded solutions of > 100 mg 1-l of high molecular weight humic matter, which after deionization and freeze-drying could only be redissolved by addition of NaOH to pH 6.

L.__! 0 Carbonyl Groups + Phrnollc Group8

6 OOC-UFa

S-FAb

DOG-HF’

V

S-HAb

Fig. 1. Dist~bution of acidic functional groups in soil fulacids (S-FA), soil humic acids (S-HA), unfractionated organic matter .from . _ river water _ (DOGUF), and^ the high * molecular weight fraction 01 organic matter trom river water (DOC-HF). ‘BECK et al. (1974), REUTERet al. (in preparation). bS~~~~~~~~ and KHAN (1972), and references cited therein. vie

and GOH (1971), river water humics belong to spectral type II, which is characteristic of low molecular weight soil fulvic acids, whereas the i.r. spectrum of the high molecular weight GPC fraction is more adequately described as spectral type I, to which most humic acids belong. Number average molecular weights (M,,) determined by vapor pressure osmometry on both unfractionated river water humic matter and its GPC fractions range from 3095 to 528 for the latter, with a,, = 1269 for the unfractionated material. This value is higher than those reported for soil fulvic acids by SCHNI~ZERand SKINNER (1968) and HANSEN and SCHNITZER(1969), which are 688 and 951, respectively. It is perhaps this specific analytical datum which distinguishes between river water humic matter and soil fulvic acids. The latter result from a more defined extraction-fractionation process, whereas aquatic humic substances probably originate by several processes (see below), of which soil extraction by meteoric waters is only one contributing parameter. From the above discussed analytical data it also becomes clear that aquatic humics contain high molecular weight components similar to humic acids. The relative size of this high molecular weight fraction is rather difficult to determine since collection, filtration, extraction and purification procedures most severely affect this particular fraction. In rivers with high DOC concentrations such as the dark colored coastaf plain streams of the SE. United States, the high molecular weight humic matter is probably the source of the Aocculant humic aggregates that can be seen to accumulate along the

ABIDANCE

OF ORGANIC

CARBON IN

RIVER WATER The Mississippi River, whose drainage basin covers about 47% of the aggregate drainage area leading to the shores of the United States, delivers about 37% of the corresponding stream discharge. The waters of the Mississippi River are derived from drainage subsystems of great climatic diversity and a corresponding variety of soils. Therefore, the water chemistry of the Mississippi River main stem in southern Louisiana can be considered the result of a large-scale geochemical averaging process. With respect to organic carbon compounds, another import~t aspect has to be considered: downstream waters of the Mississippi River have carried organic substances from far distant sources. Therefore, biopolymers (see discussion above) with short half-lives have mostly been eliminated by microbial assimilation. REUTER et d. (in preparation) found indications of rapid disappearance of dissolved polysaccharides downstream from the source area of the dissolved river water organic matter. Thus, at least 80% of the organic matter in lower Mississippi waters can be expected to consist of humic substances. MALCOLM and Dv~m (1976) have reported average DOC and POC values of 3.4 f 0.6 and 3.8 & 1.6 mg l-i, respectively, for the lower Mississippi River in water year 1969. These data correspond to an average TOC value of 7.2 & 2.2 mg l-l, which is in excellent agreement with TOC values of 8.3 _t 6.0 mg 1-i in water year 1974 and 6.2 + 1.8 mg 1-l in water year 1975 (U.S. Geol. Suru., Water Resources Data for Louisiana, 1974, 1975). Assuming that about 80% of the TOC in the lower Mississippi River occurs as dissolved humic substances, the concentration of the latter amounts to ~tc13 and 2: 10 mg 1-i in water year 1974 and 1975, respectively. Assuming further an average molecular weight of 2: 1000 (see above), one arrives at an average molar concentration of

J. H. REUTERand E. M. PERDUE

328

“v 10-5 moles 1-l for dissolved humic substances in lower Mississippi River water. Total organic carbon values both higher and lower than the above averages have been reported for different river systems. MAIERet al. (1974) give TOC values for upper Mississippi River and confluents (St. Croix and Minnesota River) ranging from 9 to 33 mg l-l, with a mean of 22 mg l-“. These high values are comparable to those reported for rivers draining the coastal plains in the southeastern U.S. (BECK et al., 1974; RFXTERet al., in preparation; U.S. Geol. Sure., Water Resources Data jar Georgia, 1973, 1974; MALCOLMand DURUM, 1976) and are often found in regions with abundant river swamps. Measured concentrations of humic substances range as high as 100 mg I-‘, or -lob4 moles l-l, with mean values of about 45 mg l-l, or -y5 x lo-’ moles 1-i. Another important North American river, the Columbia River (draining lo’/, of the aggregate drainage area of the United States and delivering N 167; of the corresponding total river discharge) has considerably lower TOC con~ntrations, ranging from 2.1 to 4.4 mg l-i, with a mean of 3.0 mg I-’ (U.S. Geof. Sure., Water Resources Data for Oregon, 1974). Compared to the Mississippi River system, the Columbia has twice the discharge per unit area. Also, a large fraction of the drainage basin is covered by mountain soils and desert soils with low organic matter content. Both factors contribute to the observed low TOC concentrations, corresponding to estimated molar concentrations of humic substances of ~5 x 10m6 moles 1-l.

ORIGIN

OF HUMIC SU~TANCES RIVER WATER

IN

The striking similarity between aquatic humics and soil humic substances suggests a soil origin for at least part of the aquatic humic matter. Similar observations have been made by MALCOLMand DURUM (1976). However, shallow ground waters generally exhibit very low organic carbon concentrations ranging from 0.1 to 1.3 mg l-‘, with a mean of 0.8 mg 1-l (LEENHEER et al., 1974) which are about one order of magnitude lower than those found in rivers. Vertical transport of dissolved humic substances in soils is generally restricted downward to the B horizon. Observations on some southeastern United States coastal plain streams show that during prolonged periods of low discharge the organic carbon concentrations decrease. On the other hand, sharp increases in discharge are usually correlated with increases in organic carbon concentration. This suggests that the bulk of the organic carbon, and specifically of the humic substances, is derived from the sail after the water level has risen upward through the profile during rain storms and overland flow has become a significant fraction of the total runoff. The data from the Mississippi River do not show any correlation between discharge and organic carbon concentration.

This observation is significant, inasmuch as it shows that large increases in discharge do not lead to a dilution effect regarding organic carbon, which is another indication that the bulk of organic carbon compounds originates with the runoff and that in situ synthesis, e.g. by primary productivity, can only play a subordinate role. An additional mechanism for the contribution of humic substances to river waters through in sifu synthesis merits discussion. Secondary eflluents from waste water treatment still retain considerable concentrations of organic matter, the bulk of which is in the form of humic substances (REBHUN and MANKA,1971; MANKAet al., 1974). These humics are apparently synthesized in the raw wastes and during passage of the wastewaters through trickling filters and activated sludge processes by browning reactions. These reactions involve the condensation of biomonomers (e.g. amino acids, sugars) and transformation into rearranged products of high reactivity. These initial products degrade by dehydration and t&ion into a variety of reactive compounds which polymerize to brown nitrogeneous polymers and copolymers (HODGE, 1953). HOERING (1973) has found these browning products (melanoidins) to be very similar to soil humic substances with regard to elemental composition, oxygen containing functional groups, and i.r. absorption spectra. Ever increasing amounts of river and ground waters are diverted for human use, loaded with organic compounds (the bulk of which will be humic substances after treatment) and returned to the rivers. Rivers such as the Ohio River, with large urban and industri~ centers, experience this diversion, use, and loading several times. Under these conditions the man-made contribution of refractory dissolved humic substances to river waters can become significant. ACII%BASE EQUILIBRIUM CHARACTERISTICS OF AQUATIC HUMIC SUBSTANCES

Inasmuch as the major fraction of dissolved organic matter in most natural waters closely resembles soil fulvic acids, it should be possible to use the extensive information available on metal-fulvic acid interactions to approximately describe those between metals and dissolved humic substances in natural waters. Much of the literature concerning metal-fulvic acid interactions published prior to 1970 has been examined in extensive reviews by FLAIG et ai. (1975) and by SCHNITZERand KHAN (1972). Relevant to any analysis of metal-fulvic acid interactions is a knowledge of the acid-base equilibrium characteristics of fulvic acid. Recently, several theoretical and experimental procedures for determination of the nature and quantity of acidic functional groups in fulvic acids have been proposed. GAMBLE(1970, 1972) using a modification of SIMMS’(1926) original treatment of titration data for polyprotic acids, has considered fulvic acid as an irregular polymer of

Importance of heavy metal-organic matter moderately low molecular weight with a number of chemically non-identical acidic functional groups whose respective dissociation ‘constants’ are a function of the overall degree of ionization of the fulvic acid polymer. Using Gamble’s method, it is possible to calculate an overall acid dissociation function for any point in the titration. More importantly, the acid dissociation functions for the ionization of individual functional groups can also be calculated. While the overall procedure is quite straightforward, the usual difficulties of locating acid-base equivalence points in the titration of humic substances (e.g. VAN DIJK, 1960; POSNER, 1964) were also encountered by Gamble. Gamble reported the presence of at least two types of acidic functional groups, including a highly acidic ‘Type I’ carboxyl group which was believed to be an o-hydroxybenzoic acid group. Other researchers (e.g. STEVENSON et al., 1973) have been unable to establish evidence for the existence of these highly acidic carboxy1 groups. However, it should be pointed out that Gamble’s evidence for the existence of Type I carboxy1 groups is based primarily on conductivity titrations, in which a minimum conductance was observed at a much lower pH than the equivalence point of an analogous potentiometric titration. This same phenomenon has been observed by other researchers (CHATTERJEE and BOSE,1952; HALLA and RUSTON,1955; PIRET et al., 1960; DUNN, 1974) and has been attributed to such diverse phenomena as ion-pairing and aggregate formation. It is currently impossible to estimate the relative contribution of each of these phenomena toward reducing the conductivity during the titration of humic substances, so the existence of Type I carboxyl groups is still somewhat speculative. Furthermore, since low pK, values (where K, is the acid dissociation constant) are common for o-substituted benzoic acids (o-OH, pK, = 2.97; o-COOH, pK, = 2.95), the assumption that Type I carboxyl groups are o-hydroxybenzoic acids has also not been substantiated. Despite the uncertainty concerning the nature and quantity of Type I carboxyl groups, the total carboxyl content of humic substances can be determined from potentiometric titrations (e.g. POMMERand BREGER, 1960; GAMBLE,1970, 1972), particularly if Gran functions (GRAN, 1950, 1952) are used to locate the titration end point. If only the total carboxyl content is known, it is still possible to use a modified version of Gamble’s model to obtain overall dissociation functions and individual group dissociation functions for fulvic acid. It is expected that the general versatility and applicability of Gamble’s model will probably lead to its more widespread usage in the characterization of the acidic functional groups of humic substances. One difficulty often encountered during the titration of humic substances is the apparently sluggish manner in which acid-base equilibrium is established. To overcome this problem, rather tedious ‘discontinuous titration’ procedures have often been used

329

(POMMERand BREGER,1960; SCHNITZERand DESJARDINS, 1962; SCHNITZERand GUPTA, 1965). On the other hand, BORGCAARD (1974a) has reported that air oxidation apparently modifies the acidic functional groups of humic acid. Therefore, he recommended that titrations of humic substances should be carried out as quickly as a normal acid-base titration. The analysis of titration data obtained by this procedure (BORGGAARD, 1974b) indicated the presence of 1.5-3.0 mequiv g 1 of highly acidic (pK, = 2.8-3.4) carboxyl groups, 2.74.2 mequiv g-’ of moderately acidic (pK, = 4.9-5.1) carboxyl groups, and 1.1-1.7 mequiv g -I of weakly acidic (pK3 = 9.4-9.7) phenolic hydroxyl groups. The great amount of information thus obtained suggests that the standard acid-base titration may indeed be very useful in characterizing the acidic functional groups of humic substances. NATURE OF METAL-ORGANIC INTERACTIONS OF AQUATIC HUMIC SUBSTANCES In recent years, a number of traditional methods such as continuous variations (JOB, 1928), ionexchange (MARTELLand CALVIN,1952), and potentiometric titrations (BJERRUM,1941) have been used to study metal-fulvic acid interations. SCHNITZERand HANSEN(1970), who have used both the ion-exchange equilibrium method and the method of continuous variations, have recommended the latter method for the study of metal-fulvic acid complexes. They also pointed out that stability constants vary considerably with pH and ionic strength. The sensitivity of /&values to experimental conditions could possibly account for the range of stability constants for a single metalfulvic acid complex which are often encountered in the literature (e.g. CHEAM,1973). GAMBLE’S(1970, 1972) description of the acidic functional groups of fulvic acid has been extended (GAMBLEet a/., 1970) to encompass metal-fulvic acid interactions. Using the ion-exchange method, Gamble et al. have calculated chelation equilibrium ‘functions’ for the formation of Cu(IItfulvic acid complexes. The chelation ‘functions’ did not vary in any systematic manner with the degree of ionization of ‘Type I’ carboxy1 groups. ARDAKANIand STEVENSON (1972) have re-evaluated the ion-exchange method of MARTELLand CALVIN (1952), and have shown that the method is applicable to metal-fulvic acid systems only under limited conditions. A modified method of data treatment and analysis was proposed which eliminates many of the sources of error in the ion-exchange method. This system was verified with known metal-organic complexes and was then used to study the Zn(II)-humic acid system. Potentiometric titration methods were used by VAN DIJK (1971) and by STEVENSON et al. (1973) to study metal-humic acid complexes. STEVENSON et al. (1973) suggested that the stability constant (R) for the reaction between a metal ion and an undissociated com-

330

J. H. REUTER and E. M. PERDUE

plexing site would be less affected by the degree of ionization of the humic polymer than would the stability constant (fl) for the reaction between a metal ion and a dissociated complexing site. The following reactions illustrate this point:

fulvic acid complexes for all these metals. These workers used the acid ~ssffiiation ‘functions’of GAM-

(1970) together with their ex~~rnen~l potentiometric titration data to calculate free metal ion concentrations. The calculated values were in excellent agreement with free metal ion concentrations deterp = (MA) mined experimentally with specific ion electrodes. M++A-*MA (M+)(A-) These results support the general applicability of GAMBLE’S (1970, 1972) method for characterization of ~MA)(H+) M++HA=MA-t-H+ iT= acidic functional groups in humic substances. How(M+)(HA) ever, MANNINGand RAMAMOORMY (1973) believe that HA = H+ + AE# _ (H+)(A-), the complexing site is a phthalic acid group rather (HA) than a salicylic acid group, as has been suggested by SCHNITZERand HANSEN(1970) and by GAMBLEet al. (1970). It is known that ga decreases with increasing Gel permeation chromatography has recently been degree of ionization of the humic polymer (GAMBLE, applied to the study of metal-fulvic acid interactions 1970, 1972). Furthermore, p values should increase (HENDRICKSON et al., 1974; MANTOURAand RILEY, with increasing degree of ionization of the polymer, 1975). Unlike most of the previously used techniques, since /l increases with increasing pH (SCHNITZERand where low pH and high ionic strength were required, HANSEN,1970). The two effects should approximately cancel, leaving E relatively inde~ndent of the degree stability constants may be determine by gel permeation chromatography under conditions which closely of ionization of the humic polymer. This predicted behaviour of K is supported by the results of GAMBLE reflect natural freshwater systems. For example, MANTOURA and RILEY (1975) have shown that at pH 8, et al. (1970), who reported that chelation equilibrium ‘functions’ (analogous to K) were insensitive to the the stability constant of the Cu(II)-fulvic acid complex is several orders of magnitude larger than at pH degree of ionization of ‘Type I’ carboxyl groups, Recently, Mets-fuivic acid interactions have been 3 (SCHNITZERand HANSEN,1970), indicating that the studied using a specific ion electrode to measure di- degree of metal-fulvic acid complex formation in natural waters is much greater than was previously rectly the activity of free metal ion in metal-fulvic acid solutions. CHEAM (1973)and CHEAM and GAMBLE expected. By studying the concentration dependence (1974) have thus determined stability constants for of the complexation of Cu(I1) by fulvic acid, MANfulvic acid complexes with Cu(II), Hg(II), Cd(I1). TOURA and RILEY(1975) have also shown that Cu(I1) C&AM (1973) reported that the stability constant for is bound to fulvic acid through two different binding the Cu(II)-fulvic acid complex is a function of the sites. The relative abundances of the binding sites and mole fraction of Cu(II), particularly if Xcu < 0.3. the respective stability constants for the binding of Much of the apparent discrepancy in reported K Cu(I1) to each site were calculated. While most studies of metal-fulvic acid interactions values for the Cu(II)-fulvic acid system was shown have concentrated on characterization of dissolved to result from the dependence of K on Xc-. GARDINER(1974) and DUNN (1974) have used speci- complexes, it should be pointed out that other fic ion electrodes to study Cd(II~hum~c substance in- mechanisms of interaction between humic substances teractions. GARDINER(1974) showed that humic suband metal ions may be important. For example, stances bind Cd(I1) to a greater extent than any of RASHID(1974) and GUY et al. (1975) have shown that the major inorganic ligands (CO:-, SO:-, Cl-, trace metals are readily adsorbed on the surface of OH-) in natural waters, particularly where humic particulate humic substances. Since a small fraction hgands are predominant. Furthermore, the degree of of the total organic carbon is particulate in most natural waters, the contribution of this fraction to complexation of Cd(H) by humic acid increases substantially with increasing pH. metal transport cannot be neglected. GAMBLE(1973) using specific ion electrodes, has The ability of humic substances and related model demonstrated that fulvic acid binds both Na+ and compounds to reduce Fe(II1) has been investigated K*. The ability of humic substances to bind alkali by SZ~LAGYI(1971, 1973) and by THEIS and SINGER metal cations must undoubtedly affect the binding of (1973, 1974). The residual concentration of Fe(I1) other metals (e.g. transition metals), p~ticularly when which is often found even in well-oxygenate waters is due not only to the formation of stable Fe(H)sodium or potassium salts are used as supporting electrolytes. The magnitude of this effect has not yet humic substance complexes but also to the reduction been determined. of Fe(II1) to Fe(I1) by humic substances. MANNING and RAMAM~ORTHY (1973) and RAMAEXTENT OF METAL-ORGANIC INTERACTIONS MOORTIIY and MANNING(1974) have studied the interOF AQUATIC HUMIC SUBSTANCES actions of fulvic acid with Cu(II), Pb(II), Cd(H), and Zn(II). They report that mixed fulvate-phosphate It is apparent from the preceding discussion that complexes are more important than simple metalsoil humic substances readily form complexes with BLE

Importance of heavy metal-organic matter trace metals in carefully controlled laboratory studies. However, in natural waters where the concentration of dissolved organic matter may be quite low and where competition between trace metals and major cations (e.g. Ca *+, Mg*+) for available complexing sites may be significant, the importance of metalfulvic acid complexes has been questioned (STUMM and MORGAN,1970). The effect of Ca*+ competition on the complexation of Cu(I1) by fulvic acid was examined by GAMBLEand SCHNITZER(1973). At a fulvie acid concentration of 1 mg l- ’ and at a (Ca)/(Cu) ratio of 80, they calculated that the fulvic acid chelating sites would be almost fully loaded with Ca*‘, with very little Cu(I1) complexation. Since fulvic acid concentrations in natural waters are often much greater than 1 mg l-‘, we have extended the calculations of GAMBLEand SCHNITZER (1973) to include fulvic acid concentrations ranging from 1 to 100 mg 1-l (3 x 10e6-3 x 10m4 M) and a range of Cu(I1) concentrations (1 x 10-*-l x 10e5 M). A Ca*+ concentration of 4 x low4 M and an initial pH of 5.0 were used in all calculations, along with the chelation equilibrium functions for the Ca*+-fulvic acid complex (R = 1.8) and the Cu(II)fulvic acid complex (i? = 23) which were reported by GAMBLEand SCHNITZER(1974). The results which are given in Fig. 2 predict that the degree of complexation of Cu(I1) and Ca* +, as well as the fraction of occupied complexing sites in fulvic acid, are essentially independent of the analytical concentration of Cu(I1). However, these parameters are quite dependent on the concentration of fulvic acid. For example, when 1 mg 1-l of fulvic acid is present, only about 8% of Cu(I1) is complexed; however, at a fulvic acid concentration of 10 mg l-‘, almost half of the Cu(I1) will be complexed. In view of the earlier discussion of the abundance of dissolved humic substances in stream waters, a fulvic acid concentration of 10 mg Fulvlc

I

Acid

(rngll

1

IO

100

“, 0.8X 0 cr E 0.8~ 0”

log

(Complrxing

Sites)

Fig. 2. Degree of complexation of Cu(I1) and Ca” as a function of fulvic acid concentration. Total Cu(I1) concentrations 1 x lo-’ M (-) and 1 x 10d5 M (---).

331

1-l is probably more representative than a concentration of 1 mg 1-l. Thus, the data of GAMBLEand SCHNITZER(1973) predict that substantial quantities of Cu(I1) could be complexed by fulvic acid even in the presence of a 40,000-fold excess of Ca*+. Experimental evidence for competition among metals for available complexing sites in natural waters has been observed by BECKet al. (1974), who reported that both Fe and Al show moderately good linear correlations with dissolved organic matter in the Satilla River in southeast Georgia. Furthermore, since an even better correlation exists between dissolved organic matter and the sum of Fe and Al concentrations (PERDUE et al., 1976), it seems likely that the lack of a good correlation between dissolved organic matter and Fe which is often observed in natural waters (LAMAR,1968) is due to competition between Fe and other metals (e.g. Al, Ca) for available complexing sites. Recently, several methods have been used to estimate the apparent complexing capacity of natural waters (G;~CHTERet al., 1973; HANCK and DILLARD, 1973; CHAU et al., 1974; RAMAMOORTHYand KUSHNER,1975). All these methods measure the concentration of complexing sites which are either unoccupied or occupied by easily displaced metals. Since most of the complexing sites are occupied by strongly bound metals such as Fe(II1) or AI(III), the apparent complexing capacity represents only a small fraction of the total complexing capacity of a water sample (CHAU et al., 1974). For example, while a 7 mg 1-l solution of fulvic acid (see above discussion) contains approximately 20 pmole 1-l of potential complexing sites (GAMEZLE et al., 1970), apparent complexing capacities are typically &2 pmole 1-l. It should be pointed out that calculations which are based on the stability constants of metal-fulvic acid complexes can only approximately describe metal-organic interactions in natural waters. Since fulvic acid is defined as the fraction of soil organic matter which remains in solution when an alkaline soil extract is acidified to pH 1, the chemical composition of such material cannot be expected to be identical to that of organic matter which has been mobilized from soils by meteoric waters. Furthermore, stability constants for metal-fulvic acid complexes have generally been determined at 25°C in low pH, high ionic strength solutions. As MANTOURAand RILEY (1975) have shown, such stability constants differ significantly from those determined under more environmentally realistic conditions. Nevertheless, if due consideration is given to the limitations of these stability constants, they can provide a qualitative or semiquantitative description of metal-organic interactions in natural waters. SUMMARY AND CONCLUSIONS The foregoing discussions have shown the following: 1. the bulk of dissolved organic carbon compounds

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in natural waters consists of biopolymers (e.g. polypeptides, polysaccharides) and geopolymers (humic substances). Since the chemical character of the geopolymers requires specialized, less common enzyme systems for breakdown, microbial attack introduces a strong bias favoring digestion of the biopolymers. As a result, the latter make up only a small fraction of the dissolved carbon, expecially in river water where primary productivity is low. Humic substances are therefore the main contributors to dissolved organic carbon. 2. The main fraction of the dissolved humic substances in river waters closely resembles soil fulvic acids. An indeterminate, yet small fraction (
aquatic humics is dependent on the concentration of humic substances and the competition for available complexing sites between trace metals and major cations. Assuming an average humic concentration of 10 mg I-’ for river waters, and using chelation stability data derived from soil humic substances, it can be shown that complexation of trace metals can be significant even in the presence of high excesses of major cations.

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