contaminant-mediated photobleaching of wetland chromophoric dissolved organic matter

EnvironmentalScienceProcesses & Impacts

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Contaminant-me

aEnvironmental Science Graduate Program,

43210, USAbDepartment of Civil, Environmental, and

University, Columbus, OH 43210, USAcSchool of Earth Sciences, The Ohio State

E-mail: [email protected]; Tel: +1-6

Cite this: Environ. Sci.: ProcessesImpacts, 2014, 16, 2098

Received 3rd March 2014Accepted 17th April 2014

DOI: 10.1039/c4em00138a

rsc.li/process-impacts

2098 | Environ. Sci.: Processes Impacts

diated photobleaching of wetlandchromophoric dissolved organic matter

Maureen C. Langlois,a Linda K. Weaversb and Yu-Ping Chin*c

Photolytic transformation of organic contaminants in wetlands can bemediated by chromophoric dissolved

organic matter (CDOM), which in turn can lose its reactivity from photobleaching. We collected water from

a small agricultural wetland (Ohio), Kawai Nui Marsh (Hawaii), the Everglades (Florida), and Okefenokee

Swamp (Georgia) to assess the effect of photobleaching on the photofate of two herbicides, acetochlor

and isoproturon. Analyte-spiked water samples were irradiated using a solar simulator and monitored for

changes in CDOM light absorbance and dissolved oxygen. Photobleaching did not significantly impact

the indirect photolysis rates of either herbicide over 24 hours of irradiation. Surprisingly, the opposite

effect was observed with isoproturon, which accelerated DOM photobleaching. This phenomenon was

more pronounced in higher-CDOM waters, and we believe that the redox pathway between triplet-state

CDOM and isoproturon may be responsible for our observations. By contrast, acetochlor indirect

photolysis was dependent on reaction with the hydroxyl radical and did not accelerate photobleaching

of wetland water as much as isoproturon. Finally, herbicide indirect photolysis rate constants did not

correlate strongly to any one chemical or optical property of the sampled waters.

Environmental impact

This contribution examines the role of herbicides on the alteration of chromophoric dissolved organic matter (CDOM) by sunlight in wetland surface waters. Ourpaper is the rst to show that organic compounds, which undergo oxidation by excited CDOM triplets can also catalyse the destruction of chromophores capableof absorbing photons found in sunlight. We also demonstrated that conventional photolysis experiments conducted in sealed phototubes in simulated ornatural sunlight resulted in the depletion of dissolved oxygen within the reactor. Thus, this second phenomenon may have ramications with respect to howphotochemical experiments are designed and executed as well as the data that is generated.

Introduction

Wetlands have the potential to attenuate mobile contaminantstransported by runoff associated with agriculture. Relativelylong retention times (compared to streams) and a favorableaspect ratio with respect to sunlight exposure make wetlandswell suited as photoreactors for contaminant degradation viadirect or indirect photolysis. Direct phototransformation oforganic contaminants can occur in wetland waters, but many ofthese substances lack the chromophores capable of absorbingphotons present in sunlight. A more likely pathway occurs whensunlight activates other wetland water constituents (photosen-sitizers) that in turn react with contaminants via indirectphotolytic pathways. These photosensitizing species includenitrate/nitrite, iron, and chromophoric dissolved organic

The Ohio State University, Columbus, OH

Geodetic Engineering, The Ohio State

University, Columbus, OH 43210, USA.

14-292-6953

, 2014, 16, 2098–2107

matter (CDOM), which lead to the formation of reactive oxygenspecies (ROS) such as the hydroxyl radical (OHc), singlet oxygen(1O2), and superoxide (O2

�/HO2�) as well as other excited

species including triplet-excited state dissolved organic matter(3DOM) and organic radicals. These species are unstable andquickly react with other water constituents including DOM(which includes CDOM and non-chromophoric molecules),carbonate species, water, and contaminant molecules.

Photobleaching, i.e., the alteration of CDOM composition byboth direct and indirect photolysis, is an important physico-chemical process in wetland chemistry and contributes to thecycling of organic matter in watersheds. For this reason, CDOMis depleted in sunlit waters more rapidly than DOM ingeneral.1–4 Over time, sunlight exposure decreases the averagesize of DOM molecules, measured both in terms of lowermolecular weight and changes in light absorbance.5 In wetlandsthe processing of DOM into smaller components has beenattributed more to photolytic thanmicrobial processes6 and hasbeen shown to affect the subsequent bioavailability of DOM toheterotrophic microorganisms.7–9

The extent to which CDOM photobleaching inuences thetransformation of organic contaminants depends on the

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Paper Environmental Science: Processes & Impacts

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chemical and optical properties of the wetland in addition to itshydrology, location, andmorphology. Attempts have beenmadeto model these processes using the chemical10–12 and optical13,14

properties of the surface waters, but much is still unknownabout the responsible pathways, particularly in wetlands.15–18

Finally, recent research has demonstrated that DOM can inhibitthe transformation of contaminants known to react readily with3DOM.19,20 One implication of these studies is the possibilitythat contaminant molecules can act as reactive species capableof transforming CDOM in irradiated waters, which to ourknowledge has not previously been explored.

For this study we investigated (1) the role of various wetlandwater chemical constituents and DOM optical characteristics onthe photodegradation of two herbicides isoproturon (IP) andacetochlor (AC), and (2) the link between contaminant degra-dation and the coincident CDOM transformation. Both of ourtarget compounds are pre-emergent broadleaf herbicides usedin large-scale agriculture and they have chemical properties thatmake them suitable for use in laboratory studies (low volatility,negligible hydrolysis rates at ambient temperatures, etc.). Theobjectives of the current study, therefore, are (1) to quantify therates of change of CDOM and of the probe compound herbi-cides AC and IP during the course of solar irradiation in variouswetland waters, (2) to relate these changes to water chemistryand optical characteristics, and (3) to determine the extent towhich light-induced changes in CDOM and in the herbicidesare interconnected.

Materials and methodsSite descriptions

Water was collected from four freshwater wetland sites in theU.S. between June 2010 and January 2011 to capture the greatestdiversity of natural photosensitizers. Yocom Farm wetland,Champaign County, Ohio (YCF) is a small (<1 ha) modiedforested wetland receiving tile drainage from corn and soybeanelds in central Ohio. Kawai Nui Marsh (KNM) is a restoredfreshwater marsh on Oahu, Hawaii (335 ha). The EvergladesWildlife Management Area Water Conservation Area 2B (EVG),located near Ft. Lauderdale, FL, is positioned at the upstreamend of the Okeechobee/Everglades system that covers much ofsouthern Florida. Samples were taken from a channel emergingfrom an expansive stretch of vegetated marsh (USGS EDENstation S11A-H, upstream). The Okefenokee Swamp (OKS) is a1.77 � 105 ha, peat-lled “blackwater” swamp that is the sourceof the Suwannee River. Samples were taken from the edge ofdesignated openmarsh area accessed from the eastern entranceto the Okefenokee National Wildlife Refuge, Georgia. In allcases, samples were taken from the shore of inundated wetlandareas with minimal emergent plant shading. Samples werecollected with a glass container on a 3 m extender rod from adepth of 30–50 cm and were vacuum-ltered as soon as possible(<2 h) with pre-combusted Gelman A/E (1 mm) glass lters andstored in the dark on ice until being refrigerated at 4 �C. In thecase of OKS samples, high turbidity of the water necessitatedsplitting the sample and ltering only half of the sample in the


eld (split OKSFF) and the remaining half in the lab two dayslater (split OKSLF).

Water chemical characterization

Parameters measured in the eld at each sampling time includewater temperature and dissolved oxygen (YSI EcoSense DO200,YSI Environmental, Inc. Yellow Springs, OH). The pH andconductivity of the ltrate were measured immediately aerltering (Thermo Orion model 130 conductivity meter, Beverly,MA). Laboratory characterization of the samples included UV-visible (UV-Vis) absorbance (Varian Cary 13 UV-visible spectro-photometer, Agilent Technologies, Santa Clara, CA, U.S.A.),dissolved organic carbon (DOC) analysis (Shimadzu TOC-VCPN, Columbia, MD, U.S.A.), total iron (Varian Vista AX CCDICP-AES, Agilent Technologies, Santa Clara, CA, U.S.A.) andinorganic ions (Dionex DX-120, Sunnyvale, CA, U.S.A. and Ska-lar San++ Nutrient Analyzer, Breda, Netherlands).

Optical characterization of water samples

UV-visible absorbance values were corrected for the absorbanceof water (Millipore Milli-Q water purication system, Billerica,MA). Specic UV absorbance (SUVA254) values were calculated inthe manner of Weishaar et al.,21 using the equation,

SUVA254 ¼ A254/[DOC] (1)

where A254 is the absorbance of the wetland water at l¼ 254 nmmeasured in a 1 cm path-length cuvette and [DOC] is expressedin moles C per L (M) and/or mg C per L (mg-C/L). Samples werediluted if their absorbance were greater than 1 and DOC was re-measured for the diluted samples.

To calculate spectral slope, Napierian absorption coefficientsfor l ¼ 250–500 nm were calculated as follows for each water,

al ¼ (2.303Al)/l (2)

where al ¼ the absorption coefficient (m�1) at l, Al ¼ correctedUV-visible absorbance at l, and l ¼ path length (1 cm for ourinstrument). A plot of ln a as a function of l yields a relativelylinear curve whose slope (S) in various ranges were determined;here S275–295, S350–400 and SR, the ratio of S275–295/S350–400 werecalculated in the manner of Helms et al.5

Absorbance was also used to calculate a screening factor (SF)for each water sample. Briey, SF at any given l is calculated asfollows,

SF ¼ (1 � 10�Al)/(2.303Al) (3)

where the variables are the same as dened above. SF is plottedas a function of wavelength and tted to a fourth-order poly-nomial (R2 > 0.98), which is then integrated over the wave-lengths of interest and subtracted from the value without lightscreening.22

Photolysis experiments

Herbicides were obtained fromChem Service (West Chester, PA)at the highest available purity. Wetland water samples were

Environ. Sci.: Processes Impacts, 2014, 16, 2098–2107 | 2099


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adjusted to the eld-measured pH using HCl or NaOH ofhighest available grade (iron free), spiked with aqueous stocksolutions of IP or AC to form a 5 mM reaction solution, andstirred open to air at room temperature for 2–3 hours. Fulvicacid solutions containing Suwannee River or Pony Lake fulvicacids (SRFA or PLFA) were prepared at pH 8 in a 1 mM bicar-bonate solution.

Quartz photoreaction tubes (path length¼ 0.9 cm, volume¼8 mL) were lled with reaction solution and sealed with airtightcaps in the dark. Reactors were irradiated in a solar simulatortted with a collimating lens (Suntest CPS+, Atlas MTT LLC,Chicago, IL). Samples were irradiated for at least one half-lifefor both analytes. Dark controls consisted of foil-wrapped tubeswere placed concurrently within the solar simulator. Directphotolysis rates were determined by irradiating analyte solu-tions in the absence of DOM and other photosensitizers.

Overall light intensity of the solar simulator was determinedusing both p-nitroanisole/pyridine and p-nitroacetophenone/pyridine actinometry,23 which is calibrated to the knownquantum yield of the reaction at 313 and 366 nm. The lightemitted by our solar simulator (on the manufacturer's 550 Wm�2 setting) was determined to be �2.0 times as strong as thesummer sun at 40�N latitude.24 Irradiance over the course ofexperiments was continuously monitored using a UV-lightradiometer (VWR International, Bridgeport, NJ) that integratessignal over l ¼ 320–390 nm. The temperature of the solarsimulator was maintained at 27 � 3 �C.

For a subset of photolysis experiments, UV-Vis absorbanceand dissolved oxygen (DO) were measured at time-points duringirradiation. DO was determined using a micro-DO probe (LazarResearch Laboratories, Inc., Los Angeles, CA) dry-calibrated inair and wet-calibrated using air-saturated water. DO wasmeasured in most cases within 1 hour of removal from the solarsimulator and aer the tubes had equilibrated to roomtemperature. Inux of atmospheric O2 into the tubes duringmeasurement was determined to be minimal during initialreadings (acquisition time 2–3 min), but interfered withmeasurements over longer spans of time, precluding multiplereadings of each sample tube. Herbicide concentrations wereanalyzed using a Waters reversed-phase HPLC (Milford, MA)equipped with a Nova-Pak C18 column (3.9 � 150 mm). With amobile phase of 55 : 45 ACN : H2O (v/v), detection parameters

Table 1 Chemical constituent concentrations and optical characterizati

Watersample

Datecollected pH

[Fetot](mM)

[NO3/NO2](mM N)

Total [N](mM N)

[DOC(mg

YCF 07/01/10 8.9 3.6 1.55 586 7.20KNM 12/18/10 6.1 29 bdlc 119 32.2EVG 12/31/10 7.9 1.1 0.68 118 25.0OKSFF 01/01/11 3.5 4.3 2.64 73.7 52.5OKSLF 01/01/11 3.5 6.8 7.50 510 139

a Screening factor calculated over l ¼ 290–370 nm. b Spectral slope of linenm, and the ratio of the two, aer [5]. c Below detection limit (�10 ppb N¼lab-ltered, FF ¼ eld-ltered); YCF ¼ Yocom Farm water (Ohio).

2100 | Environ. Sci.: Processes Impacts, 2014, 16, 2098–2107

were as follows: for IP, l ¼ 240 nm, retention time z 4.7 min;and for AC, l ¼ 220 nm, retention time z 6.5 min.

We made amendments to the indirect photolysis experi-ments with EVG samples to elucidate reaction mechanisms. Toassess the role of OHc, t-butanol was added to EVG water spikedwith AC such that the nal concentrations were 25 mM and 5mM, respectively. Another set of mechanistic experimentsinvolved bubbling EVG with argon (1 min mL�1 solution) todisplace atmospheric gases and then preparing photolysissolutions in a glovebox lled with N2/H2 (95%/5%). Initialreadings on these low-O2 solutions showed that this treatmentyielded DO of 35–40% of atmospheric O2 saturation.

Kinetic modeling

The observed rate of loss in complex mixtures such as naturalwaters is attributed to both direct and indirect photolysis. Todetermine the rate constant that describes direct photolysisoccurring in a given wetland water (kdir), the rate constant fordirect photolysis in water was multiplied by the appropriatescreening factor (Table 1). Indirect reaction rate constants (kind)were then calculated by subtracting kdir from kobs (eqn (4)),

kind ¼ kobs � kdir (4)

where kobs is the overall rate constant from each experiment.

ResultsChemical and optical characteristics of wetland waters

The four wetlands sampled in this study represent a broadrange of chemical composition, from the relatively low-DOC,moderately alkaline YCF to the turbid, acidic OKS (Table 1). Theheterogeneity of CDOM composition precludes a simplechemical quantication of the colored fraction presumed todominate photochemical activity. Accordingly, UV-visibleabsorbance spectra for each water sample were used to indi-rectly characterize CDOM optical properties. Screening factorsfor all sampled waters aligned inversely with [DOC], with thehighest [DOC] sample (OKSLF) absorbing nearly 50% of ambientsunlight (Table 1). SUVA254 values (an estimate of aromaticity)of the wetland waters fell roughly into two categories, with YCFand EVG at �360 M�1 cm�1 (�2.95 L mg C�1 m�1) and KNMand OKSFF at �225 M�1 cm�1 (�1.85 L mg C�1 m�1). High

on of collected wetland water samples

]C/L)

SUVA254(M�1 cm�1) SF(290–370)

a S(275–295)b S(350–400)

b SRb

� 0.2 372 0.921 0.015 0.015 1.02� 0.2 221 0.780 0.012 0.015 0.81� 0.1 351 0.814 0.020 0.020 0.99� 0.7 230 0.687 0.014 0.013 1.04� 0.7 174 0.513 0.014 0.015 0.93

arized UV-vis absorbance over the ranges l ¼ 275–295 nm, l ¼ 350–4000.7 mM) EVG¼ Everglades water; OKS¼Okefenokee Swamp water (LF¼



Table 2 Rates of change in optical characteristics and dissolved oxygen of wetland waters during irradiation

Wetland waters

Rates of change of photobleaching parametersCorrelation between DO andspectral rates of changea

DOb S275–295c S350–400

c SRd S275–295 S350–400 SR

EVG Alone �1.04 � 0.3 1.08 � 0.01 �0.263 � 0.01 68.5 � 0.1 0.81 0.74 0.87AC only �1.35 � 0.3 1.34 � 0.01 �0.303 � 0.04 70.9 � 0.2 0.83 0.14 0.93IP only �1.71 � 0.1 0.94 � 0.02 �1.77 � 0.07 143 � 0.5 0.36 0.45 0.46AC, Low-O2 0 0.97 � 0.03 �0.446 � 0.04 74.1 � 0.3 0.07 0.20 0.18IP, Low-O2 0 1.09 � 0.01 �0.463 � 0.05 81.2 � 0.2 0.03 0.00 0.06AC, t-butanol �1.06 � 0.3 1.42 � 0.01 0.348 � 0.01 52.0 � 0.1 0.93 0.72 0.95

OKSFF AC only �2.96 � 1.0 1.78 � 0.02 1.07 � 0.04 43.7 � 0.1 0.74 0.77 0.55IP only �5.76 � 1.4 2.10 � 0.02 0.033 � 0.01 83.0 � 0.1 — — —

OKSLF AC only �3.74 � 1.7 1.32 � 0.02 1.14 � 0.03 15.1 � 0.1 0.70 0.80 0.08IP only �9.42 � 1.6 2.63 � 0.04 2.34 � 0.03 29.4 � 0.1 0.97 0.96 0.97

a Coefficient of determination (R2) for rates of change of spectral parameters (S) indicated as function of DO. b Rate of change of DO (�10�2% h�1)compared to 100% atmospheric saturation. c Rate of change of spectral slope (S) region indicated (�10�3 h�1). d Rate of change of spectral slope (S)ratio (� 10�2 h�1) AC ¼ acetochlor (5 mM); DO ¼ dissolved oxygen; EVG ¼ Everglades water; IP ¼ Isoproturon (5 mM); OKS ¼ Okefenokee Swampwater (LF ¼ lab-ltered, FF ¼ eld-ltered). Errors represent 95% condence intervals.


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aromaticity has been reported in the northern Evergladespreviously.25

Changes to CDOM during light exposure

During the course of irradiation experiments in the solarsimulator, EVG and OKS water samples were periodicallymonitored for UV-visible absorbance and DO. In all cases, theabsolute UV-Vis absorbance of the samples decreased with lightexposure, indicating loss of chromophores over time. This trendis consistent with numerous previous observations ofphotobleaching.1,6,8

S275–295 increased steadily with light exposure for all waterstested, implying greater loss of absorbance at the higher l

(lower energy) end of this range (Table 2). Increases in S275–295for our samples corroborate previous studies of photobleachingin natural waters and is indicative of the production of lowermolecular weight moieties.5,26 While S275–295 changed at similarrates among measured waters (mean rate of change ¼ 1.39 �10�3 h�1), the S350–400 parameter was more variable (Table 2).Helms et al.,5 also observed both phenomena for various DOMsamples with increasing S275–295 as a function of irradiance andvariability in S350–400.

Water samples spiked with AC or IP (5 mM) affected the rateof spectral slope change during the irradiation of EVG water(Table 2). Dark controls showed that the initial spectral slopesfor EVG water in the presence and absence of the herbicideswere identical, indicating that the contribution of analyte lightabsorbance to the total absorbance of the sample is negligible.Further, the presence of each herbicide caused a different effecton spectral slope with time. With AC, dS275–295/dt and dS350–400/dt changed slightly compared to no herbicide; however, theslope ratio SR remained constant. By contrast, the presence of IPconsiderably accelerated the rate at which S350–400 decreased,causing a rate of change of SR over twice that of no-herbicide orAC-containing solutions. Therefore, the presence of herbicidesalters the way in which the CDOM changes over the course of


irradiation. We believe that this has not been observed inprevious DOM mediated herbicide photodegradation studies.

Reducing the initial DO concentration to sub-oxic levels(dened here as argon purged solutions) in our experimentsslightly lowered the rate of change in S275–295, for AC spikedsolutions perhaps implying that photobleaching of chromo-phores absorbing lower-wavelength photons is more due todirect photolysis as oppose to the presence of reactive oxygenspecies. The opposite occurred for the rate of change in S275–295,for IP spiked solutions whereby we observed a slight increase.Additional experiments would be necessary to determine theeffects of low-O2 conditions in the presence of herbicides. Ourresults corroborated those reported by Gao and Zepp,27 whoobserved some photobleaching in anoxic Satilla River water.

In OKS water, all experiments were performed in the pres-ence of herbicides. Compared to irradiation with AC, for bothOKSFF and OKSLF the presence of IP resulted in more rapidabsorbance changes than those observed in EVG (Table 2). Inboth OKS splits, the presence of IP caused dSR/dt to doublerelative to AC. The factor of two increases in SR rate of changecaused by the presence of IP was consistent among EVG, OKSFFand OKSLF waters despite their disparate water characteristics.

Changes in oxygen levels during light exposure

Irradiated EVG and OKS waters were monitored for changes inDO over the course of irradiation (Table 2). O2 is a participant inmany known indirect photolysis reaction pathways includingthe photobleaching of CDOM,4,9 the formation of singletoxygen28 and photo-Fenton reactions29 such that photo-bleaching has been shown to be largely due to indirect photo-oxidation. From a starting value of 100% atmospheric satura-tion, we observed steady decreases in DO during all irradiationexperiments relative to dark controls. Further, the rate of DOloss was relatively consistent for each water sample oversuccessive trials.



Fig. 1 DO loss rate as a function of concentration of dissolved organiccarbon ([DOC]) or total iron ([Fe]) in irradiated wetland waters spikedwith either acetochlor (AC; a and c) or isoproturon (IP; b and d). Pointsrepresent average of all trials. Error bars indicate standard deviation.Samples stored in acidified conditions, but brought to field pH forexperiment were included in this regression (6 acidified out of 19 total)to have access to more points.

Fig. 2 UV-visible absorbance spectra and chemical structures for theherbicides acetochlor (AC; black line) and isoproturon (IP; gray line).

Fig. 3 Change in herbicide concentrations (a: acetochlor, AC and b:isoproturon, IP) during light exposure in Everglades (EVG) water underthree conditions: standard (equilibrated in air, D), low-O2 (B), or with25 mM t-butanol (AC only, ,). Dark symbols indicate dark controls.Error bars represent 95% confidence interval.


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Changes in spectral slope in the presence of AC or IP weremirrored by accelerated DO loss relative to wetland water in theabsence of the analytes (Table 2). With AC, DO decreased �30%faster than in wetland water alone, while changes in the pres-ence of IP, the effect was even larger (�70%). Experimentsconducted in EVG water under low-O2 conditions resulted inlittle change in DO level over the course of the irradiation,although readings exhibited a greater standard deviationcompared to air-saturated irradiations. The rate of DO loss wascorrelated to both DOC and total iron (Fetot) for the wetlandwaters, consistent with direct and Fenton-mediated photo-oxidation (Fig. 1).

For all waters, the presence of IP caused substantially (1.3 to2.5 times) higher rates of DO loss compared to AC. Thisenhancement mirrors the increased photobleaching of waterscontaining IP versus AC described above. Interestingly, the samepattern emerges in the dependency of DO loss rate on DOC andFetot shown in Fig. 1 i.e., twice the DO loss per unit DOC or Fetotoccurred for solutions containing IP versus AC.

Rates of contaminant loss during light exposure

Direct photolysis. The rates of direct photolysis of AC and IPwere measured by irradiating the herbicides in water. The directphotolysis reaction rate constant for AC in water was 7.5� 0.6�10�3 h�1 (half-life: �3.9 d) and is consistent with a previouslymeasured value (8.2 � 10�3 h�1).35 The kdir for IP was smaller at2.7 � 0.9 � 10�3 h�1 (half-life: �10.6 d). The lower kdir of IPversus AC cannot be explained by their respective absorbancespectra at wavelengths present in sunlight (Fig. 2), but isconsistent with previous observations.30


Indirect photolysis. Observed herbicide loss rate constants(kobs) were measured in irradiated wetland waters (Fig. 3) and inbuffered fulvic acid solutions. For AC, wetland water acceleratedphotodegradation over water by a factor of 2–3 even when cor-rected for light screening (Table 3). Indirect photolysis of ACproceeded with nearly identical kind for almost all the sampledwetlands (including YCF, KNM, EVG and OKSLF), with amean of1.2� 0.1� 10�2 h�1. We also observed increasing kind, for PLFAsolutions as DOC concentrations increased (7, 14 and 21 mg C/L, pH ¼ 8) whereby the range of kind is similar to thosemeasured for our wetland samples (�1.1 � 10�2 h�1, data notshown). In contrast SRFA solutions (pH ¼ 8) at low (2.5 mg-C/L)concentrations caused little or no enhancement over directphotolysis. One dramatic exception to this pattern was the eld-ltered Okefenokee sample (OKSFF) whereby kind was 4–5 timeshigher than all other waters. All dark controls showed negligible(<1%) herbicide loss.

IP photolysis was enhanced by a factor of 30–200 in wetlandwaters compared to water alone (Table 3). These results agreewith previously reported ratios of direct to sensitized photolysisfor IP.12 Similar to our observations for AC the irradiation of IPin PLFA solutions with increasing DOC yielded linearly



Table 3 Reaction rate constants for the photodegradation of AC and IP in various waters under simulated sunlight

Water sample kobsa (10�2 h�1) kdir

b (10�2 h�1) kindc (10�2 h�1)

Wetland watersd AC IP AC IP AC IP

YCF 1.9 � 0.1 12.7 � 0.4 0.69 � 0.06 0.25 � 0.09 1.2 � 0.1 12.4 � 0.8KNM 1.6 � 0.1 16.4 � 0.9 0.58 � 0.06 0.21 � 0.09 1.0 � 0.1 16.2 � 0.7EVG 1.8 � 0.1 30 � 3.0 0.61 � 0.06 0.22 � 0.09 1.2 � 0.2 30 � 2.0OKSFF 5.4 � 0.5 6.7 � 0.7 0.51 � 0.06 0.19 � 0.09 4.9 � 0.5 6.5 � 0.7OKSLF 1.7 � 0.1 4.1 � 0.4 0.38 � 0.06 0.14 � 0.09 1.3 � 0.1 3.9 � 0.4

PLFAe

7 mg C/l 1.6 � 0.1 20.6 � 1.8 0.75 � 0.06 0.27 � 0.09 0.9 � 0.1 20 � 2.014 mg C/l 2.0 � 0.2 25.5 � 3.3 0.75 � 0.06 0.27 � 0.09 1.2 � 0.2 25 � 3.021 mg C/l 2.2 � 0.1 31.9 � 3.7 0.75 � 0.06 0.27 � 0.09 1.4 � 0.1 32 � 4.0

a Observed experimental reaction rate constant. b Direct photolysis rate constant corrected for light screening by kUP$SF290–370; where kUP,AC ¼ 7.5�0.6 � 10�3 h�1; kUP,IP ¼ 2.7 � 0.9 � 10�3 h�1. c Indirect photolysis rate constant kind ¼ kobs � kdir.

d Unless otherwise noted, experimentalconditions: eld pH, atmospheric gas saturation before photolysis, 5 mM herbicide. e Fulvic acid solutions diluted from stock solution,distinguished by nal [DOC] AC ¼ acetochlor (5 mM); EVG ¼ Everglades water; IP ¼ Isoproturon (5 mM); OKS ¼ Okefenokee Swamp water (LF ¼lab-ltered, FF ¼ eld-ltered); YCF ¼ Yocom Farm water (Ohio).

Fig. 4 Indirect photolysis reaction rate constants for isoproturon (IP)as a function of dissolved organic carbon concentration ([DOC]) invarious wetland waters (filled markers; field pH, see Table 1) and fulvicacid solutions (open markers, D ¼ Suwannee River FA, , ¼ Pony LakeFA; pH ¼ 8.0). Error bars represent 95% confidence interval.


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increasing rate constants (Fig. 4), consistent with the involve-ment of 3DOM in IP photolysis as shown by others.12,30 However,as shown in Fig. 4 the correlation between PLFA concentration(as DOC), and IP kobs was not observed for our wetland waters.

Probing mechanisms of indirect photolysis

Low-O2 conditions. To determine the extent to which O2 wasinvolved in indirect photolysis, we purged solutions to about35–40% of atmosphere-equilibrated DO. For AC, low DOconditions did not signicantly alter kind, although deviationsfrom pseudo-rst order kinetics occurred (Fig. 3a). For IP, thisdeviation from pseudo-rst order behavior was even morepronounced (Fig. 3b), and the initial rate of degradation wasclearly enhanced under low-O2 conditions.

Addition of t-butanol. Indirect photolysis for AC was shutdown in EVG spiked with t-butanol, a known OHc scavenger


(Fig. 3a). The observed reaction rate constant was actually lowerthan kdir, suggesting that the quencher shut down the indirectOHc mediated pathway. DO loss in the t-butanol-spiked solu-tion was lower than an identical solution without the quencher,and more closely approximated the DO loss of EVG water alone(Table 2). Although dS275–295/dt was nearly the same as withnon-quenched AC solutions, the SR rate of change was some-what lower than even wetland water alone, suggesting someinhibition occurred in photobleaching in the presence of t-butanol (Table 2).

DiscussionEstablishing likely pathways of contaminant degradation

The herbicides used as probe compounds in this study werechosen because they are known or suspected to have differentdominant pathways for indirect photolysis. For AC, no previousdata is available for indirect photolysis in sunlit natural waters,but the primary indirect photolytic pathway for other chlor-oacetanilide herbicides including metolachlor,31,32 alachlor,16

propachlor and butachlor33 involves non-specic ROS species,particularly OHc. Our results, using the OHc quencher t-butanol(Fig. 3a), conrm that this radical is important in AC indirectphotolysis in wetland waters.

The major formation pathways of OHc in sunlit watersinclude photolysis of NO3

� and NO2� the reaction of sensitized

CDOM with H2O, and3 photo-Fenton reactions mediated byFe(III) and DOM, which becomes important under acidicconditions. Major sinks for OHc are DOM, carbonate, and, to alesser extent bicarbonate.34 Given the second-order rateconstant between AC and OHc, (kOHc ¼ 7.5 � 109 M�1 s�1),35 asteady state hydroxyl radical concentration ([OHc]ss) on theorder of 10�16 M would be required to achieve the average kind of1.2 � 0.1 � 10�2 h�1 that we observed for AC. This is a plausiblescenario given that [OHc]ss of this magnitude has been observedin various wetland waters.36




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Iron-rich, acidic, highly colored waters such as OKS areknown to undergo photo-Fenton OHc production,2,27 which mayhelp to explain why AC photolysis was three times faster inOKSFF than in any other water. Alachlor, which is structurallysimilar to AC, was shown to photodegrade much more rapidlyin acidied (pH ¼ 4) wetland waters versus unacidied,16 andthe same mechanism may be at work with AC. Although OKSLFwater was at the same pH as OKSFF, it also had nearly threetimes the DOC. Both OHc quenching and signicant inner-ltereffects associated with this high DOC may in part explain theslower AC kinetics in OKSLF. Finally, it is possible that the ironspeciation between OKSFF and OKSLF may have also beenaltered over the intervening two days, which could also affectthe reactivity of the two samples.

The primary reactive species responsible for IP trans-formation in natural waters appears to be 3DOM12,37 andpossibly DOM-derived organic radicals,38 although IP oxidationhas been observed with other reactive species including OHc.39

The increased rate of IP photolysis in low-O2 conditionsobserved by us corroborates this pathway as dissolved oxygencan compete with IP for 3DOM (Fig. 3b). We observed a linearincrease in kind with increasing concentration of PLFA (Table 3and Fig. 4) and corroborate the results reported by Gereckeet al.12 Surprising there was no linear dependence of kind onDOC (Fig. 4) for any of the wetland waters.

We observed no strong correlations between IP kind and waterproperties such as absorbance, DOC, or pH, but our resultsrevealed trends. The highest kind value occurred for EVG (pH ¼7.9) while the slowest kinetics was observed for OKS (pH ¼ 3.5).While differences in DOM composition will in part explain ourobservations these results also support previous ndingswhereby the rates of triplet-induced degradation increase withpH.37 For example Bruccoleri et al.,40 demonstrated that thequantum yield for the formation of triplet DOM increasessignicantly with increasing pH. Stosser et al.,38 reported asimilar phenomenon for the production of organic radicals. Theidea that pH is correlated with greater IP phototransformation issupported by Fig. 4, whereby the two high-pH water samples withsimilar SUVA254 values (EVG and YCF) fall close to the trendcreated by PLFA (buffered at pH 8). With respect to the inuenceof DOM composition Guerard et al.41 observed that photo-degradation of the antibiotic sulfadimethoxine (also primarilydegraded by 3DOM in sunlight) occurs most quickly in PLFAsolution and is not photosensitized by SRFA. We believe that thisis partly due to the different antioxidant properties of the fulvicacids whereby the more terrestrially derived SRFA possessesantioxidant poly-phenolic moieties that are not present in PLFA,which is derived from microbial organic matter precursors.20

Fig. 5 Indirect photolysis reaction rates (kind) for isoproturon andacetochlor in wetland water samples used in this study.

Contaminant acceleration of photobleaching with IP

DOM photobleaching by both direct and indirect photolysis hasbeen attributed to the loss of chromophores. Surprisingly, bothAC and IP inuenced photobleaching by changing spectralslope and DO beyond the photolytic rates observed in wetlandwater alone (Table 3). To date we are unaware of any previousstudy that showed the effect of pesticides on DOM


photobleaching. This suggests that the herbicides themselvesare involved in the formation of an additional pool of reactivespecies that can oxidize CDOM. For AC, the change in photo-bleaching and DO loss compared to solutions in its absence wassmall. Conversely, IP accelerated the rate of DOM photo-bleaching substantially. This effect was minimized under low-O2 conditions, suggesting involvement of DO.

We hypothesize that IP and other compounds susceptible tooxidation by triplets may form radical cations (IPc+) throughreaction with triplet aromatic ketone moieties in 3DOM asshown in eqn (5).42,43 Others have demonstrated that the triplet-sensitized oxidation of phenylureas and other aniline-contain-ing compounds, decreases in the presence of DOM.19,20 Theproposed mechanism of this inhibition is DOM-inducedreduction of IPc+ back to the parent IP (eqn (6)) as opposed tothe continued reaction of IPc+ with other unknown species (M)to form other products (denoted here as IPox).

3DOM + IP / DOMc� + IPc+ (5)

IPc+ + DOM / IP + DOMc+ (6)

IPc+ + M / IPox + Mc+ (7)

Our observation of increased photobleaching of CDOM inthe presence of IP is consistent with reduction of the radicalcation IPc+ by CDOM as shown in eqn (6). In high-DOC solutionsOKSFF and OKSLF, the presence of IP increased both spectralchanges and consumption of O2 compared to AC. However, therate of IP degradation itself is lowest in both OKS samples(Fig. 5), which is contrary to the observation that IP reaction rateroughly increases with increasing DOC.12 The fact that IPincreases photobleaching without being rapidly degradedsuggests a pathway similar to eqn (6) as opposed to eqn (7).

Our work has identied aspects of these photochemicalinteractions that merit further work in order to understand thefate of compounds that predominantly undergo oxidationthrough reactive triplets and their possible role in photo-bleaching. In particular, the role of DOM chemical compositionhas not yet been fully investigated. DOM with higher aromaticcarbon moieties is both more likely to inhibit triplet-mediatedcontaminant oxidation (due to the presence of naturally occurring




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anti-oxidant compounds present in the higher plant precursors)20

and these substances are more likely to form 1O2.14,18 Our ndingsimply that for more aromatic (allochthonous) DOM, we wouldexpect to see a greater reduction in IP kind and possibly a morerapid decrease in DO. Finally, 3DOMmoieties may have a range ofredox potentials. For example recent studies have suggested thepresence of multiple triplet pools with varying susceptibility toreduction by borohydride.18,42

Methodological implications of photochemical experimentaldesign

Changes in DO over the course of a long (8–30 h) irradiationexperiment reveal issues in the application of closed-tubephotoreaction experiments. Photochemical experiments oenuse airtight, capped phototubes (e.g., ref. 16 and 44 and othersreferenced therein). While studies have compared solar simu-lators and natural sunlight,45,46 little attention has been devotedto issues regarding the use of sealed phototubes to determinenatural water photoreactivity given the potential for artifactscaused by O2 depletion. Our observations of herbicide degra-dation kinetics in this study show that later samples deviatesomewhat from pseudo-rst order kinetics, with more parentcompound le than would be predicted by the kinetic model.This deviation is more pronounced in solutions such as OKSthat contain high [DOC] and occurs more frequently with IPthan AC, which correlated to higher rates of O2 consumption asdiscussed above. This occurred even though IP irradiationswere consistently shorter in duration than AC (IP:�8 h, AC:�25h). Whatever the source of the O2 depletion, further study of thisphenomenon is warranted, as the fundamental chemical reac-tivity of photosensitizers including 3DOM are altered in thepresence of lower DO.14,18

While we observed a strong correlation between DO deple-tion and spectral slope changes over the experimental time-frames in this study, it can be argued that low-O2 conditionsmay not be representative of shallow wetland waters (assumedto be well-mixed and aerated at the top layers where photo-reactivity is most relevant). Comparative studies between closedand open solar simulation experiments are needed in order todetermine whether differences in the mechanisms and rates ofcontaminant phototransformation exist and if they point to apreferred set of photofate testing procedures in natural waters.

Acknowledgements

This study was supported by funding from the Ohio WaterDevelopment Authority grant OWDA-4968, the National ScienceFoundation grants CBET-0504434 and CBET-1133094, and theOhio State University Environmental Science GraduateProgram. Special thanks to Dr SueWelch for nutrient analysis ofwater samples, Ross and Mary Yocom of Yocom Farm, KrispinFernandes, Clark Liu and Philip Moravcik of the University ofHawaii, Maya Wei-Haas for help with the graphics, andMatthew Yin for statistical analysis and consultation. Finally,the authors thank two anonymous reviewers, whose commentsgreatly improved this paper.


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contaminant-mediated photobleaching of wetland chromophoric dissolved organic matter

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