Environ. Sci. Technol. 2003, 37, 4190-4198
Photoformation of Low-Molecular-Weight Organic Acids from Brown Water Dissolved Organic Matter THOMAS BRINKMANN, PHILIP HO ¨ RSCH, DANIEL SARTORIUS, AND FRITZ H. FRIMMEL* Water Chemistry, Engler-Bunte-Institute, University of Karlsruhe, 76128 Karlsruhe, Germany
This work describes the effects of simulated solar UV light on the bulk properties of dissolved organic matter (DOM) of bog lake water and on the formation of lowmolecular-weight organic acids (LMWOAs). By means of sizeexclusion chromatography it was shown that the more hydrophilic moieties of the DOM were preferentially photodegraded while the more hydrophobic ones remained relatively unaffected or were even formed. The combined photochemical-biological degradation proved to be more important than the pure photochemical mineralization. Formic, acetic, pyruvic, oxalic, malonic, and succinic acids were identified as important degradation products. Their contribution to the dissolved organic carbon increased from 0.31% before to 6.4% after 24 h irradiation. About 33% of the bioavailable photoproducts of DOM were comprised of these LMWOAs. The influence of nitrate on the formation of carboxylic acids could not be observed in the investigated system. Kinetic experiments indicated that degradation of LMWOAs occurred simultaneously during irradiation experiments, R-oxygen-substituted LMWOAs being more amenable to these processes. Dissolved iron acted as a catalyst of DOM photodegradation and LMWOA photoformation. Copper played an antagonistic role in the irradiation experiments, reducing the formation of formic, acetic, and malonic acids while increasing the formation of oxalic acid.
Introduction The absorption of sunlight by dissolved organic matter (DOM) induces a multitude of photoreactions (1-3). As a consequence, DOM is bleached and partially mineralized, and its average molecular weight is reduced (3-7). A lot of lowmolecular-weight substances have been identified as photochemical breakdown products (4, 8). Among these are nutrients such as ammonium and phosphate, carbon gases such as carbon dioxide and monoxide, aliphatic carbonyl compounds, and carboxylic acids (4, 9-13). The latter are thought to result mainly from the high-molecular-weight fraction of DOM (14, 15). Many investigations have shown that irradiation of refractory DOM increased its bioavailability and therefore stimulated heterotrophic bacterial activity (4, 6, 8, 11, 1620). However, there is an increasing number of publications * Corresponding author phone: +49 721 608 2580; fax: +49 721 699154; e-mail:
[email protected]. 4190
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which report the contrary effect of DOM becoming more recalcitrant due to irradiation, or even no effect at all (8, 21-27). One key parameter seems to be the nature of the irradiated DOM. Allochthonous, humic-rich DOM generally becomes more bioavailable after photodegradation, while autochthonous, algal-derived DOM or readily bioavailable substrates become less bioavailable. The latter could be due to the formation of recalcitrant products of higher molecular weight or to the mineralization of the organic compounds (8, 21, 28). Low-molecular-weight organic acids (LMWOAs) are photochemical degradation products of DOM which are easily biodegradable (4, 11, 20). However, only a few studies have demonstrated a direct correlation of the photochemical production of LMWOAs and changes of the biological properties of DOM. For example, Kieber et al. showed that the photochemical production rate of pyruvic acid in seawater was highly correlated to its biological uptake (14). Furthermore, Bertilsson and Tranvik found that nearly all of the photochemically produced formic acid was biologically mineralized, while a significant portion of acetic and especially malonic acid was incorporated into bacterial biomass (11). The sunlight-induced photochemical and combined photochemical-biological degradations of DOM play important roles within the global biogeochemical carbon cycle. The annual loss of oceanic DOM via these processes was estimated to be 2-3% of the dissolved organic carbon (DOC) (4). Furthermore, photochemical degradation is considered to be one of the major sinks of riverine DOM in the oceans (8, 29). Despite the importance of the natural photoproduction of LMWOAs, it is not well-known if they play a significant or even key role within the biologically labile fraction of irradiated DOM. Furthermore, factors influencing their formation have, to our knowledge, only scarcely been investigated. No influence of iron on the photoproduction of LMWOAs from DOM of 38 different lakes was found by Bertilsson and Tranvik (30). In a recently published work, Goldstone et al. (31) showed that the reaction of hydroxyl radicals with humic substances led to the production of LMWOAs. However, the formation rates of hydroxyl radicals in natural environments were calculated to be too low to explain the experimentally observed concentrations of LMWOAs. Dissolved iron is known to act as an effective photocatalyst for DOM oxidation in aquatic systems. Following absorption of sunlight, aqueous iron complexes can undergo a ligand to metal charge transfer, which reduces Fe(III) to Fe(II). In the case of inorganic iron-hydroxo complexes such as Fe(OH)2+, the very reactive hydroxyl radicals (HO•) are formed, and in the case of organic iron-carboxylate complexes, it leads to the formation of carboxylic acid radicals (DOMCO2•), which subsequently may undergo decarboxylation. Iron(II) is rapidly reoxidized to Fe(III) in aerated waters by different oxygen species. This results in the formation of superoxide radicals, hydrogen peroxide, or hydroxyl radicals, all of which are reactive oxygen species capable of further oxidizing DOM (2, 32-35). It is yet unknown if this redox cycle also induces the formation of bioavailable lowmolecular-weight organic compounds (36). In the case of LMWOAs, their presumable production from DOM should be accompanied by their iron-catalyzed photolysis (37-39). Similarly to the redox reactions of Fe(III), complexes of Cu(II) can be reduced by sunlight to Cu(I) while the ligand is oxidized. The primary step is followed by a dynamic redox cycle, where superoxide radicals, hydrogen peroxide, and 10.1021/es0263339 CCC: $25.00
2003 American Chemical Society Published on Web 08/14/2003
TABLE 1. Basic Sample Characteristics of the Drain Water of Lake Hohloh HO18ba F(DOC) κ pH a254nm a436nm
20.6 ( 0.5 mg/L 31 µS/cm (19 °C) 4.3 (20 °C) 111 m-1 7.6 m-1
c(Fe) c(Cu) F(NO3-) F(SO42-)
6.2 ( 0.5 µmol/L 0.25 ( 0.01 µmol/L 1.41 ( 0.02 mg/L 2.5 ( 0.1 mg/L
a κ ) electrical conductivity, and a 254nm and a436nm ) absorption coefficients at 254 and 436 nm. Some data are given with 95% error intervals.
hydroxyl radicals are formed (40-42). However, copperDOM complexes show lower absorptions than the corresponding iron compounds (40). In addition, paramagnetic metal ions and especially Cu(II) are known to quench the luminescence of DOM and thus excited states, which are prerequisites for potential photochemical reactions (43). Some recent publications have shown the inhibiting effect of Cu(II) in photodegradation experiments (44, 45). The objective of the presented work was to investigate specific factors such as nitrate, irradiation time, dissolved iron, and dissolved copper influencing the release of biologically labile LMWOAs upon solar irradiation of DOM. In addition, the relative contribution of these carboxylic acids to the bioavailability of DOM was studied. Changes in the bulk properties of DOM were characterized by size-exclusion chromatography (SEC). SEC coupled to on-line DOC detection is a powerful tool for the characterization of DOM (46). The separation is mainly based on hydrodynamic radii. Small molecules are able of permeating into the column gel and are retarded. For comparison purposes, the separation is often described by the dimensionless distribution coefficient Kd, which is calculated from the retention times of the analytes, tr, of totally excluded molecules, te, and of totally permeating molecules, tp (47):
Kd )
tr - te tp - te
(1)
For totally excluded molecules, Kd becomes 0, and for totally permeating ones, it becomes 1. The separation based on size is always accompanied by specific interactions of the analytes with the column material. For example, deprotonated carboxylic acid groups are repulsed by negatively charged groups on the gel surface. As a result, the retention time and the Kd value decrease. On the other hand, nonpolar aliphatic and aromatic moieties may exhibit strong hydrophobic interactions with the gel, which often lead to Kd > 1 (47, 48).
Experimental Section Materials. Deionized water (18.2 MΩ‚cm, Milli-Q PLUS, Millipore) was used in all experiments. All chemicals used were commercially available with purities g99.0% except DFOM (deferoxamine mesylate; Aldrich, ∼95%). The DOM sample HO18b was taken in May 2000 from the surface of a bog lake drain (Lake Hohloh, Black Forest, Germany). Subsequently, the sample was filtered (0.45 µm, cellulose nitrate) and stored in the dark at 4 °C. Basic sample characteristics are listed in Table 1. Analytical Methods. The UV/vis absorption was recorded on a Cary 50 spectrophotometer (Varian) with an optical path length of 1 cm and background corrected. The resolution was equal to 1 nm. Total concentrations of dissolved metals were measured by ICP-OES (inductively coupled plasma optical emission spectrometry) on a Vista-Pro CCD simultaneous ICP-OES
spectrometer (Varian) at 234.350 nm (iron) and 324.754 nm (copper). Yttrium (F(Y) ) 2 mg/L) served as internal standard at 377.433 nm. Size-exclusion chromatography was performed according to a previously described method using a TSK HW 40S gel column (25 cm × 2 cm i.d.; eluent 1.5 g/L Na2HPO4‚2H2O, 2.5 g/L KH2PO4, pH 6.6; flow rate 1 mL/min; injection volume 2 mL) and on-line UV and DOC detection (46). The ionic strength of the samples was adjusted to that of the eluent by addition of a suitable volume of a solution containing the phosphate salts of the eluent in the appropriate higher concentrations. A cylindrical thin film reactor with a rotating inner cylinder and a low-pressure mercury lamp in the middle served as DOC detector. DOM was oxidized to CO2, which in turn was quantified by a nondispersive IR spectrometer (49). For the calculation of Kd, t0 was determined using blue dextran. tp was determined by injecting deionized water and using an electrical conductivity detector. DOC values alone were measured by operating the system in the bypass mode, i.e., without chromatographic separation. LMWOAs, nitrate, and sulfate were quantified by ion exchange chromatography on a DX 500 chromatographic system (Dionex, Sunnyvale, CA), which was configured with polyether ether ketone (PEEK) material and equipped with a GP40 gradient pump, an EG40 eluent generator, and an AS40 autosampler. Degassed water (helium 5.0) was used for production of the potassium hydroxide eluent in the eluent generator EG40. The anion exchange column AS 11 (250 mm × 4 mm i.d.) and the guard column AG 11 (50 mm × 4 mm i.d.) were used for separating the analytes at room temperature. A volume of 0.7 mL of sample was flushed through the six-port rheodyne injection valve, the actual injection volume being set to 100 µL. Anions were separated at a flow rate of 2.0 mL/min with an aqueous potassium hydroxide eluent and a gradient program (0.0-5.0 min isocratic, c(KOH) ) 0.20 mmol/L; 5.0-15.0 min linear gradient, 0.20-15.0 mmol/ L; 15.0-20.0 min linear gradient, 15.0-21.7 mmol/L; 20.130.0 min regeneration, 35.0 mmol/L; 30.1-40.0 min equilibration, 0.20 mmol/L). The detection system consisted of an electrical conductivity detector (ED40, cell temperature 35 °C, temperature compensation 1.7%/°C) and an anion selfregenerating suppressor (ASRS I) run in the autosuppression recycle mode at 300 mA. All the operations and the integration of the signals were performed with the PeakNet 5.10 software (Dionex). Data were stored at an acquisition rate of 5.0 Hz. Relative standard deviations of six replicate injections (c ) 10 µmol/L) were below 4%, and method detection limits were below 0.2 µmol/L. The identity of the LMWOAs was confirmed by capillary zone electrophoresis with indirect UV detection on an HP 3DCE system (Hewlett-Packard) according to a modified method of Soga and Ross (50). In contrast to the original method, water was first injected by pressure (50.0 mbar, 4.0 s) to achieve electrostacking conditions. Following this, the sample was introduced by electrokinetic injection (-5 kV, 45 s). Biodegradation experiments were performed with unirradiated and irradiated solutions according to Hambsch et al. (51). First, 450 mL of the irradiated sample was diluted by addition of 225 mL of deionized water to have sufficient volume for the different analyses. A 400 mL sample of the resulting solution was laced with 40 mL of a solution containing inorganic nutrients to render organic carbon the only growth-limiting factor. After sterile filtration (0.22 µm, mixed cellulose ester), 5 drops of a solution containing a natural biocenosis obtained from Lake Hohloh was added, and the samples were incubated at room temperature in the dark for 100 h under stirring and air saturation. The inoculum had been previously obtained by filtration of 5 L of the original sample (0.45 µm, polycarbonate), resuspension of the filter VOL. 37, NO. 18, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 2. Concentrations of DOC, BDOC, and CDOC of Lake Hohloh Water HO18b before and after Irradiation and Biodegradationa
a
sample
G(DOC), mg/L
unirradiated before biodegradation unirradiated after biodegradation irradiated before biodegradation irradiated after biodegradation
11.0 ( 0.04 10.4 ( 0.07 10.4 ( 0.03 8.46 ( 0.05
G(BDOC), mg/L
G(CDOC), mg/L
G(CDOC)/G(DOC), %
10.2 10.2 8.11 7.30
92 98 78 86
0.64 2.0
DOC values are given with standard deviations.
in 20 mL of sterile NaCl solution (0.9%), and filtration of the suspension with a glass fiber filter. SEC experiments were done with both irradiated and unirradiated samples. The samples denoted “before biodegradation” were filtered sterile (0.22 µm, mixed cellulose ester) immediately after addition of the inoculum and stored at -17 °C until further analysis. The samples denoted “after biodegradation” were filtered after the 100 h incubation experiment (0.22 µm, mixed cellulose ester) and subsequently analyzed. The measured DOC values of the samples were corrected for the different dilution steps. The data in Table 2 correspond to the DOC concentration of the solution that was used in the irradiation experiment. Irradiation Experiments. Irradiation experiments were carried out between August 2000 and September 2001. Before irradiation, different solutions were added to the original sample (sodium nitrate, iron(III) chloride, DFOM, copper(II) sulfate; hydrochloric acid (c ) 1 mol/L) and sodium hydroxide (c ) 1 mol/L) for pH adjustment) to investigate the influence of specific parameters. The dilution factor remained constant within one experimental series. In the case of DFOM and metal ions, samples were allowed to equilibrate for one week at room temperature in the dark. All glassware used for irradiated samples was heated for 24 h at 110 °C to obtain sterile conditions and to prevent biodegradation of the LMWOAs formed. Prior to irradiation, samples were filtered sterile (0.22 µm, mixed cellulose ester); afterward, they were stored at -17 °C and subjected to the various analytical methods within 2 weeks. During the time of irradiation, the unirradiated samples were stored at room temperature in the dark. Irradiation experiments were done with stirred, airsaturated samples at 22 ( 2 °C. Nine samples could be irradiated simultaneously (for every series eight DOM samples and one blank sample containing deionized water), the surface area being O ) 10.8 cm2, the volume V ) 35.6 mL, and the optical path length d ) 3.3 cm. For the investigations on biodegradability, a single sample of a bigger volume was irradiated (O ) 145 cm2, V ) 479 mL, d ) 3.3 cm). The commercially available solar UV simulator (Oriel) consisted of an ozone-free 1000 W xenon arc lamp, the spectrum of which was fitted to natural solar UV light by an atmospheric attenuation filter (Schott WG 320) and an additional 6 mm filter plate (Schott WG 295). The spectrum of the solar UV simulator contained relatively small portions of visible light, while the cutoff wavelength in the UV range was situated at 292 nm. The radiant flux densities measured by spectral radiometry were 6.01 W/m2 for 292-320 nm (UV-B), 102 W/m2 for 292-400 nm (UV), and 150 W/m2 for 292-500 nm. A more detailed description of the experimental setup was given by Frimmel (6). The photon flux density measured by polychromatic ferrioxalate actinometry (52) within the wavelength range of 290-500 nm accounted for PP ) 4.67 × 10-4 einstein/(m2‚s). To compare samples with different absorptions, for example, induced by differing pH values or DOC concentrations, the amount of absorbed light energy Eabs(t) for a given irradiation time t was calculated by using 4192
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∫
Eabs(t) ) Ot WP,λ(1 - 10-Ah λ(t)) dλ
(2)
λ
where O stands for the surface area, λ for the wavelength, WP,λ for the spectral radiant flux density of the solar UV simulator (W/(m2‚nm)), and A h λ(t) for the arithmetic mean value of absorption during this time interval.
Results and Discussion Effects of Irradiation on the Bulk Properties of DOM. The investigated drain water from Lake Hohloh HO18b is acidic and highly absorbing, contains low concentrations of inorganic salts (Table 1), and is rich in humic-derived DOM (53). To study the influence of irradiation and biodegradation on the bulk properties, HO18b was diluted, the pH adjusted to 4.2, and the resulting sample irradiated for 24 h with simulated solar UV light. For the unirradiated sample before biodegradation, a single peak with a maximum at a retention time of tr ) 28.3 min was found in the size-exclusion chromatogram with DOC detection (Figure 1). Most of the eluting compounds showed values of 0.0-1.0 for the dimensionless distribution coefficient Kd. This indicated that the separation was mainly based on size (hydrodynamic radii) and that ionexclusion and hydrophobic interactions of the analytes with the column material played a minor role (47). Regarding the UV254 nm detection (Figure 2), the peak was narrower and its maximum was found at a slightly lower retention time (tr ) 27.8 min), which was well in agreement with literature results (54, 55). This means that DOM fractions with bigger hydrodynamic radii show higher specific UV absorptions. No changes were found in the chromatogram after the sample had been subjected to biodegradation with a natural biocenosis for 100 h at room temperature (Figure 1). In contrast, irradiation with simulated solar UV light led to considerable changes of the DOM properties. The maximum of the main peak shifted to tr ) 29.7 min, which was indicative of a reduction of the mean hydrodynamic radii and presumably of the mean molecular weight too. New fractions appeared in the chromatogram within the retention time windows of tr ) 42-48 min and tr ) 36-39 min. In SEC experiments with standard compounds, we found that these retention times were typical for aliphatic mono- and dicarboxylic acids, for example, formic and oxalic acids. The new fractions became smaller after incubation with the biocenosis, indicating that a major part of the photoproducts was biologically labile. However, the irradiated sample after biodegradation (dotted fat line) still showed higher DOC signals for tr > 33 min than the unirradiated sample before biodegradation (straight thin line). We therefore assume that the irradiation of DOM also yielded biologically stable photoproducts. These compounds were UV-absorbing, as could be seen from the SEC-UV254nm results (Figure 2). Irradiation of DOM led to a decrease of F(DOC) from 11.0 to 10.4 mg/L (Table 2). On the other hand, the bioavailable dissolved organic carbon (BDOC), which resulted from the difference of F(DOC) before and after incubation, increased from 0.64 to 2.0 mg/L. Such an increase in bioavailability after irradiation is typical for humic-rich, refractory DOM (21). Compared to the pure photochemical mineralization
TABLE 3. Concentrations of LMWOA before and after Irradiation and Biodegradation of Lake Hohloh Water HO18b ∑(carboxylic acids)
FIGURE 1. Size-exclusion chromatograms with DOC detection of Lake Hohloh water HO18b before and after irradiation with simulated solar UV light for 24 h (pHt)0 4.2, pHt)24h 4.3) as well as before and after biodegradation using a natural biocenosis (t ) 100 h). The samples were diluted with a concentrated eluent (3:4) prior to SEC measurement. For DOC values see Table 2.
FIGURE 2. Size-exclusion chromatograms with UV254nm detection of Lake Hohloh water HO18b before and after irradiation with simulated solar UV light for 24 h (pHt)0 4.2, pHt)24h 4.3) as well as before and after biodegradation using a natural biocenosis (t ) 100 h). The samples were diluted with a concentrated eluent (3:4) prior to SEC measurement. For DOC values see Table 2. of DOM, the combined photochemical-biological degradation was more effective in removing DOC. The small portion of the DOC degraded by the biocenosis in the unirradiated sample was probably due to the addition of the inoculum, which might have been accompanied by readily bioavailable substrates. Another reason might be residual amounts of nonrefractory organic substances, which could have been degraded by the adapted biocenosis. This process would have been supported by the addition of inorganic nutrients. Additional information can be obtained from the chromatograms by calculating the CDOC (chromatographable dissolved organic carbon), which was done by integration between tr ) 20 min and tr ) 60 min (Table 2). The CDOC represents those compounds of the DOM pool which elute within this specific retention time window. The remaining DOM molecules show higher retention times and thus stronger sorption to the column material. A much stronger decrease of the CDOC as compared to the DOC was found as a result of irradiation. While the CDOC accounted for 92%
sample
G(C), mg/L
unirradiated before biodegradation unirradiated after biodegradation irradiated before biodegradation irradiated after biodegradation
0.034 0.012 0.66 0.015
G(C)/ G(C)/ G(DOC), G(BDOC), % % 0.31 0.12 6.4 0.18
3.5 33
of the DOC in the unirradiated sample before biodegradation, this value was reduced to 78% upon irradiation. This indicated a qualitative change in the DOM composition, either caused by a loss of carboxylic acid groups, leading to decreased electrostatic repulsions with the column material, or caused by increased relative proportions of nonpolar, hydrophobic moieties, rendering stronger hydrophobic interactions. The loss of carboxylic acid groups could be rationalized by decarboxylation reactions, which, for example, occur during photolysis of iron carboxylate complexes (34). The increased relative amounts of hydrophobic substances could, on the other hand, be due to a preferential degradation of hydrophilic constituents (13) or to the formation of hydrophobic coupling products (21, 28), for example, diphenyl ethers (56). The latter are typically found after irradiation of lignin model compounds (57, 58). Our results thus indicate that more hydrophilic moieties of DOM were preferentially degraded by simulated UV sunlight while more hydrophobic ones remained relatively unaffected or were even formed. The new fractions in the chromatograms pointed to LMWOAs, and indeed, formic, acetic, pyruvic, oxalic, malonic, and succinic acids were identified as photoproducts by capillary electrophoresis and quantified by ion exchange chromatography. Before irradiation, the sum of these acids was equal to F(C) ) 0.034 mg/L (Table 3). Afterward, they represented F(C) ) 0.66 mg/L, comprising 6.4% of the DOC. Biodegradation caused a significant reduction of the LMWOAs. Here, the important contribution of LMWOAs to the biologically labile fraction of the DOC could be shown, since they accounted for 33% of the bioavailable organic carbon. Hence, LMWOAs are major constituents of the biologically labile photoproducts of DOM. When related to the loss of CDOC caused by biodegradation (0.81 mg/L), LMWOAs accounted for 80%, which means that the differences observed in the size-exclusion chromatograms could be almost exclusively attributed to this class of compounds. Apart from the six LMWOAs investigated in this study, it is reasonable to assume that additional compounds were formed during irradiation. These probably comprise nonbioavailable, UV-absorbing substances as indicated by Figures 1 and 2. Furthermore, the formation of amino acids (59), carbonyl compounds, and other LMWOAs (4, 12) has already been reported in the literature. Influence of Nitrate. It is well-known that the presence of nitrate in fresh surface waters can, by forming hydroxyl radicals, induce the oxidation of DOM (35) and trace contaminants such as pesticides (2). To study the influence of nitrate, concentrations up to F(NO3-) ) 64 mg/L were added to the bog lake water, and the samples were irradiated for 24 h (pHt)0 ) 3.9). The change in the amount of light absorbed by DOM due to the inner filter effect of nitrate could be neglected because nitrate shows only a small absorption in the UV-B range (302nm/Mw(NO3-) ) 0.012 L/(mg‚m) (2)) as compared to that of the DOM (a302nm/F(DOC) ) 3.1 L/(mg‚m)), and in addition, most of the radiation absorbed by DOM was located in the UV-A range. The concentrations of the photochemically produced LMWOAs are shown in Figure 3. VOL. 37, NO. 18, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 4. Absorbed Energies by HO18b and pH Values for Different Irradiation Times and pH Starting Values G0(DOC) ) 10.2 mg/L
G0(DOC) ) 10.3 mg/L
t, h
pH
Eabs, kJ
pH
Eabs, kJ
0 2.95 6.57 7.55 13.5 18.9 24.0
4.0 4.2 4.3 4.2 4.4 4.8 4.8
0 0.928 2.06 2.35 4.09 5.81 7.16
8.0 6.3 6.6 7.0 7.0 7.1 7.2
0 1.05 2.32 2.66 4.66 6.52 8.20
FIGURE 3. Influence of nitrate on the photochemical production of LMWOAs from HO18b (G0(DOC) ) 10.7 mg/L, pHt)0 3.9, pHt)24h 4.0, t ) 24 h). For data interpretation, two factors interfering with the LMWOA measurements have to be considered. First, samples can easily be contaminated by organic acids originating from air, filters, syringes, or other laboratory material (60). Although all materials were thoroughly rinsed with deionized water before coming into contact with sample solutions, contaminations could not be avoided completely, because the samples were in contact with the ambient air during irradiation to ensure air saturation. For example, the unirradiated blank sample in this experimental series contained 0.19 µmol/L acetic acid and 2.1 kJ, we assumed that photoproduction of oxalic and pyruvic acids was still occurring. Their steadystate concentrations thus indicate that their photodecomposition was faster than that of the other LMWOAs. The common feature of lactic, pyruvic, and oxalic acids is that they bear an additional oxygen atom in the R-position to the carboxylic acid group. It is well-known that iron complexes of R-keto- and R-hydroxycarboxylic acids are photochemically more reactive than those of the corresponding carboxylic acids without R-oxygen substituents (37-39). This led us to the hypothesis that these acids were photodegraded under the catalytic influence of iron, which was present in rather high concentrations in the brown water (Table 1). Additional evidence for the influence of iron on the photodegradation process is given in the next section of this work. Influence of Iron. The original sample contained 6.2 ( 0.5 µmol/L total Fe, which is equivalent to 0.35 ( 0.03 mg/L (Table 1). This concentration is much higher than the solubility of inorganic iron(III) hydroxides at a pH value of 4.3, which is lower than 0.1 µmol/L (35). The presence of iron-DOM colloids could be ruled out since these would be totally excluded in SEC and elute at Kd values of 0 (Figure 1) (61). We therefore assumed that most of the dissolved iron was complexed by DOM and that only a minor part was present as Fe(OH)2+, which would be the dominating species in organic-free solutions (35). To investigate the influence of iron, we added FeCl3 at different concentration levels to the bog lake water prior to irradiation. Additional samples were laced with the natural complexing agent DFOM (c ) 30 and 70 µmol/L), which forms very strong, nonphotoreactive complexes with Fe(III) (32). The conditions for the 24 h irradiation experiment are given in Table 6. The size-exclusion chromatograms of the samples with higher iron concentrations were similar to those obtained in the previous experiment. For example, the sample to which 5 µmol/L iron had been added showed a single peak, which became smaller and shifted to higher retention times after 24 h of irradiation (Figure 6 as compared to Figure 1). The appearance of two new fractions was also observed. Given similar values for pH and absorbed light energy (Table 6), these changes were less pronounced in the case of the DFOM VOL. 37, NO. 18, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 6. Influence of addition of 5 µmol/L iron(III) chloride and of 30 µmol/L DFOM on the size-exclusion chromatographic behavior of Lake Hohloh water before and after irradiation (t ) 24 h). The samples were diluted with a concentrated eluent (1:4) prior to SEC measurement. For data see Table 6. (30 µmol/L)-containing sample, indicating a diminished reduction of the mean hydrodynamic radii (Figure 6). Another difference is that the retention time maximum of the second new fraction eluting between tr ) 43 min and tr ) 48 min was shifted to a higher value as compared to the sample to which 5 µmol/L iron had been added (tr ) 41-45 min). This fraction typically contains monocarboxylic acids. Their retention is strongly dependent on the ionic strength of the sample (48). In this work, the ionic strength of the sample was adjusted to that of the eluent by addition of a more concentrated eluent solution. This procedure works well because of the low electrical conductivity of HO18b. However, in the case of prior addition of the complexing agent DFOM, the resulting ionic strength of the sample was higher than that of the eluent because of the added methanesulfonate anion. This led to increased retention times for monocarboxylic acids and to the appearance of this fraction in the unirradiated sample. The effects of changes in ionic strength on retention times have already been described by Specht and Frimmel (48). The peak at tr ) 57.7 min was assigned to DFOM. The distribution coefficient was well above 1.0, which led us to conclude that hydrophobic interactions with the column material had taken place. The peak did not show any changes upon treatment of the sample with simulated solar UV light. Therefore, DFOM is unlikely to have been subjected to photochemical degradation, either directly or via reactive species produced from DOM. The influence of DFOM could be mainly attributed to its iron-complexing properties, while the scavenging of reactive intermediates seemed to be negligible. We thus conclude that the iron-catalyzed oxidation is an important pathway of the solar-induced degradation of DOM in Lake Hohloh. Additional evidence was gained from the concentrations of photoproduced LMWOAs (Figure 7). On one hand, the addition of DFOM led to a decreased production of formic and malonic acids while the formation of acetic and oxalic acids remained unaffected within the experimental error range. On the other hand, the samples with higher iron content showed an increased formation of formic, acetic, and malonic acids. It thus seems that dissolved iron is an important factor in the photoproduction of LMWOAs. However, one has to consider the fact that the addition of iron(III) chloride caused a partial precipitation of the DOM. Prior to irradiation, the samples were filtered sterile, which 4196
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FIGURE 7. Concentrations of LMWOAs (left y axis) released from HO18b upon irradiation with simulated solar UV light (t ) 24 h) as well as iron concentrations and absorbed energies per DOC (right y axis) (LOD ) limit of detection).
removed the precipitates. This is the reason why the DOC values of the unirradiated solutions decreased with increasing amounts of added iron (Table 6). Furthermore, the measured iron concentrations were lower than the ones calculated from the sum of iron of the sample without additives and the laced solutions. The UV/vis absorption of the solutions and thus the amount of absorbed energy were influenced by two effects: First, iron itself absorbed light, and second, the ironinduced precipitation of DOM caused a reduction of absorption. For the sample to which 5 µmol/L iron had been added, the first effect predominated; for higher concentrations, the second one predominated. As a consequence, the samples with higher iron concentrations absorbed lower amounts of light energy (Table 6), but because of decreasing DOC values, the amount of absorbed energy per DOC increased slightly (Figure 7). The observed increase in LMWOA photoformation might therefore be ascribed to the differences in absorbed energy. In addition, it is possible that the less photoreactive fractions of DOM were preferentially removed during precipitation. The remaining, more reactive fraction would thus have given higher concentrations of LMWOAs upon irradiation. Nevertheless, a considerable change in LMWOA formation was already observed for the samples to which the lowest amounts of dissolved iron had been added (+ 5 and + 10 µM Fe(III)) and which were least affected by precipitation (Table 6). We therefore conclude that the presence of dissolved, photoreactive iron has a stimulating effect on the photoformation of certain LMWOAs. This finding is not in accordance with the results of Bertilsson and Tranvik (30), who did not find any correlation between the natural iron content of the DOM-containing waters and LMWOA production. However, additional effects may have overshadowed the influence of iron since they investigated samples from 38 different lakes. As LMWOAs are also supposed to undergo iron-catalyzed photooxidation (37-39), the release of these acids from Lake Hohloh DOM must have been faster than their photochemical degradation. For the samples named “no additives”, “+ 5 µmol/L Fe(III)”, and “+ 10 µmol/L Fe(III)”, the ironstimulated photoformation was more pronounced for formic and acetic acids than for malonic acid, while no stimulation at all was found for oxalic acid. This result might be explained by both increasing thermodynamic stabilities of complexes of these acids with iron(III) and increasing quantum yields of their photodegradation in the order formic/acetic acid > malonic acid > oxalic acid (37, 38, 62). Thus, although the addition of iron might have stimulated the photoformation
FIGURE 8. Concentrations of formic and acetic acids released upon 24 h of irradiation depending on the concentration of dissolved copper (G0(DOC) ) 16.9 mg/L; pHt)0 5.1). of oxalic acid, the effect could not be measured because its photodegradation was stimulated to a similar extent. Further investigations should focus on the mechanisms of LMWOA formation from DOM and the influence of iron on a molecular level. LMWOAs are formed either directly from the radicals originating from the ligand to metal charge transfer or indirectly via reactive oxygen species (superoxide anions, hydroxyl radicals) produced by the iron redox cycle or by DOM itself and their reaction with DOM. Influence of Copper. The original lake water contained 0.22 ( 0.01 µmol/L Cu (16 ( 1 µg/L; Table 1). Similar to the dissolved iron, we assumed that the major part of the dissolved copper is complexed by DOM, first because copper is known to form the thermodynamically most stable complexes among the doubly charged metal ions (IrvingWilliams order, 35), and second because copper complexation capacities of Lake Hohloh DOM obtained by differential pulse polarography typically range from c(Cu)/F(DOC) ) 0.9 µmol/ mg to c(Cu)/F(DOC) ) 3.5 µmol/mg (53). In the case of the water used in this study (F(DOC) ) 20.6 mg/L), this would correspond to concentrations of complexed copper in the range of 19-72 µmol/L, which are well above the concentration in the original sample. To evaluate if the catalyzing or inhibiting effect of copper ions predominates in the sunlight-induced degradation of DOM, we added different concentrations of CuSO4 to the bog lake water. Because the nonirradiated samples did not differ significantly in their absorption properties, the concentrations of LMWOAs did not have to be normalized to the amount of absorbed energy. We found that the addition of copper slowed the photoformation of formic, acetic, and especially malonic acids (Figures 8 and 9). The inhibitory effect became significant for c(Cu) g 0.6 µmol/L. The maximum inhibition effect was found for c(Cu) ) 10 µmol/ L. Further addition of copper did not affect the production of these LMWOAs within the experimental error range. Surprisingly, the opposite behavior was found for oxalic acid, for which the formation increased up to c(Cu) ) 40 µmol/L, and from then on stayed constant. In the case of oxalic and malonic acids, the results could be confirmed in a second experiment, while in the case of formic and acetic acids, the data were not considered due to high blank values. The total amounts of the released oxalic and malonic acids differed slightly between experiments 1 and 2, because the xenon lamp used for irradiation had been replaced in the meantime. It could be argued that the lower concentrations found for formic, acetic, and malonic acids were due to the formation of stable copper-carboxylate complexes, which could not be detected by the employed ion chromatographic method. However, copper-carboxylate complexes are ki-
FIGURE 9. Concentrations of oxalic and malonic acids released upon 24 h of irradiation depending on the concentration of dissolved copper (G0(DOC) ) 16.9 mg/L for experiment 1 and 16.6 mg/L for experiment 2; pHt)0 5.1). netically very labile (63), and recovery experiments (standards with c(Cu) ) 90 µmol/L and c(LMWOA) ) 1 and 10 µmol/L) did not show any losses of LMWOAs within experimental error. Therefore, copper mainly acts as an inhibitor of LMWOA photoformation, possibly by quenching excited states (43). The border concentration above which no more inhibition was found corresponded quite well to the typical copper complexation capacities of Lake Hohloh DOM. This indicates that the binding sites for copper are also responsible for the inhibition of LMWOA photoformation. Unfortunately, this does not explain the completely different behavior of oxalic acid. Interestingly, the concentrations of oxalic acid in the unirradiated samples also increased with increasing copper concentrations, which was not due to contaminations of the reagents. The data for oxalic acid in Figure 9 were corrected for those found in the unirradiated samples. One explanation for the observed increase could be that copper displaced some of the iron from oxalate, thus decreasing the photodegradation rate of the very reactive ferrioxalate complex. However, copper oxalate complexes are thermodynamically less stable than iron oxalate complexes (62). In addition, it has been reported that Cu(II) increases the quantum yield of formation of Fe(II) from the ferrioxalate complex by increasing the rate of decomposition of the oxalate radical C2O4•- (64). To conclude, the antagonistic effect of copper(II) regarding the photoproduction of readily bioavailable LMWOAs from humic matter asks for a more detailed investigation. It would be a promising approach to investigate the influence of copper on the photodegradation of substances of known structure.
Acknowledgments This work was funded by the DFG (Deutsche Forschungsgemeinschaft) within the ROSIG project and the DVGW (Deutsche Vereinigung des Gas- und Wasserfaches). We thank Gabi Kolliopoulos and Ulrich Reichert for performing the size-exclusion chromatography and Reinhard Sembritzki for doing the metal ion analyses. We also acknowledge the critical comments of two anonymous reviewers, which helped to improve the manuscript. Parts of this work were presented at the 11th biennial meeting of the International Humic Substances Society (IHSS) in Boston, MA, July 22-26, 2002. T.B. gratefully appreciates the travel bursary award given by the IHSS.
Supporting Information Available One table. This material is available free of charge via the Internet at http://pubs.acs.org. VOL. 37, NO. 18, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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Received for review November 15, 2002. Revised manuscript received July 3, 2003. Accepted July 3, 2003. ES0263339
