Reactive Transport of Iomeprol during Stream-Groundwater Interactions

Article pubs.acs.org/est

Reactive Transport of Iomeprol during Stream-Groundwater Interactions Irina Engelhardt,*,†,# Henning Prommer,‡,§,∥ Manoj Schulz,⊥ Jan Vanderborght,† Christoph Schüth,#,▽ and Thomas A. Ternes⊥ †

Forschungszentrum Jülich, Institute of Bio- and Geosciences, Agrosphere (IBG-3), Germany CSIRO Land and Water, Western Australia § University of Western Australia, School of Earth and Environment, Western Australia ∥ National Centre for Groundwater Research and Training, Flinders University, Adelaide, GPO Box 2100, SA 5001, Australia ⊥ Federal Institute of Hydrology (BfG), Germany # Technical University of Darmstadt, Institute of Applied Geosciences, Germany ▽ IWW Rheinisch-Westfälisches Institut für Wasserforschung, Mühlheim a. d. Ruhr, Germany ‡

S Supporting Information *

ABSTRACT: The transport and biochemical transformations of the iodinated X-ray contrast medium (ICM) iomeprol were studied at the stream/groundwater interface. During a one-month field experiment piezometric pressure heads, temperatures, and concentrations of redoxsensitive species, iomeprol and 15 of its transformation products (TPs) were collected in stream- and groundwater. The data set was analyzed and transformation processes and rates identified by comparing conservative and reactive transport simulations. ICM and TP transformations were simulated as a cometabolic process during organic carbon degradation. Using iomeprol/ TPs ratios as calibration constrain mitigated the uncertainties associated with the high variability of the ICM wastewater discharge into the investigated stream. The study provides evidence that biodegradation of ICM occurs at the field-scale also for predominantly denitrifying conditions. Under these anaerobically dominated field conditions shortest simulated half-life (21 days) was in the same range as previously reported laboratory-determined half-lives for aerobic conditions.

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with respect to the field-scale biotransformation behavior of ICMs after their emission to groundwater systems,6 in particular at the stream/groundwater interface. For pharmaceutically active compounds Lewandowski et al.8 suggested that an increased activity of microorganisms can enhance the biotransformation of pharmaceuticals at this interface with significant potential for a complete elimination of wastewaterrelated micropollutants.9 While laboratory experiments allow for a better control of initial and boundary conditions, in situ experiments provide the most valuable data for the determination of the field-relevant biogeochemical transformation pathways and rates. However, to quantify transformation rates in situ, concentration changes induced by reactive processes need to be distinguished from advective-dispersive and diffusive transport. Under the complex and highly dynamic flow conditions that typically prevail at the

INTRODUCTION In many densely populated regions river and groundwater systems have become impacted by micropollutants that are released from wastewater treatment plants (WWTPs). Among these micropollutants polar and more persistent organic substances such as iodinated X-ray contrast media (ICMs), for example, diatrizoic acid, iohexol, iopamidol, and iopromide, which are used for imaging soft tissues, internal organs, and blood vessels, have become a particular concern. In 2006, global consumption of ICMs was 3.5 × 106 kg/a with a further increasing trend.1,2 ICMs are metabolically stable in the human body, and are excreted completely within a day.3 They are primarily introduced into the aquatic environment via domestic and hospital wastewaters. ICMs undergo degradation or transformation during the wastewater treatment process but are still commonly detected at elevated concentrations in wastewater-receiving rivers and groundwater.1−5 However, for yet unknown reasons, degradation varies between zero (no degradation) and 80%.6 Although the behavior of ICMs in WWTPs and drinking water treatment plants (DWTPs) has been investigated in batch experiments,7 little information exists © 2013 American Chemical Society

Received: Revised: Accepted: Published: 199

July 18, 2013 November 16, 2013 November 25, 2013 November 25, 2013 dx.doi.org/10.1021/es403194r | Environ. Sci. Technol. 2014, 48, 199−207

Environmental Science & Technology

Article

After collection water samples were analyzed in a flowthrough cell for pH and electric conductivity (EC). Dissolved oxygen (DO) was measured with oxygen-optodes (PreSens Precision Sensing GmbH, GER). Samples for analysis of metals, iomeprol and iomeprol TPs were filtered, preserved at pH 2 by acidification with 3.5 M H2SO4, and stored cooled at 7 °C until analysis. Filtered samples for dissolved organic carbon (DOC) analysis were stored frozen. Water samples were analyzed for DOC with infrared spectroscopy (Dima-TOC 100, Dimatec Analysentechnik GmbH, Essen GER). Major ion concentrations were determined by ion chromatography (IC-Dionex). Metal ions (Fe and Mn) were analyzed by atomic adsorption spectroscopy (AAS-Jena Analytik), and HCO3 by titration with HCl. Iomeprol and its 15 TPs were analyzed with LC-tandemMS as described by Kormos et al.13. Total sediment-bound organic carbon (SOC) of the aquifer material and stream bed sediments was measured after combustion with infrared spectroscopy (CHNS vario Macro CHNS, Elementar, Hanau, GER). Reactive Transport Modeling. All data collected during the study period were analyzed by a suite of flow, conservative transport and reactive transport models. The numerical models MODFLOW,14 MT3DMS,15 and PHT3D16 were used for the respective tasks. The present work builds upon a previous study in which a conservative multispecies transport model was set up and calibrated through joined inversion of measured hydraulic heads, temperatures and concentration data for the artificial sweetener acesulfame.12 A summary of the groundwater flow and conservative transport model setup and simulations is provided in the SI. The experimentally detected behavior of iomeprol13 was translated into a reaction network for PHT3D (SI Figure SI2). All equilibrium-based speciation of major ions and redox reactions were directly adopted from the original PHREEQC17 database. In addition the temperature-dependent mineralization of DOC and SOC were incorporated into the reaction database, together with the transformation kinetics of iomeprol and its daughter products. Only these newly added kinetic reactions are discussed below. The mean distribution coefficient koc of iomeprol was experimentally determined with 1.89 and the sorption affinity of iomeprol was therefore expected to be limited.18 DOC/SOC Degradation. The oxidation kinetics of DOC and the low biodegradable fraction of SOC (SOC,l) were adopted from earlier, comparable studies,19−21 and according to the Michaelis−Menten kinetic the bacterial consumption of the organic carbon was assumed to be limited by electron acceptor availability:

groundwater/surface water interface this requires (i) a comprehensive characterization of the study site, (ii) a properly designed monitoring campaign, (iii) highly sensitive and precise analytical methods, and (iv) detailed investigations with coupled numerical models that simultaneously consider the highly transient flow, transport and biogeochemical processes and how they interact. Addressing the gap in understanding the environmental fate of ICMs at the interface of surface water-groundwater systems, the goal of this study was to identify the in situ physical and chemical behavior of iomeprol under complex field-scale conditions. The ICM iomeprol was selected because of its high consumption and as it can mineralize into iodo-acid disinfection byproducts (DBP) during drinking water disinfection by strong oxidants such as chlorine or chloramines and that are reported to be genotoxic and cytotoxic in mammalian cells.10 As part of the present study we explored a range of potentially suitable modeling approaches to identify and quantify the transport and kinetic transformation of iomeprol under intense groundwater/surface interactions. The approaches were used to describe and analyze the fate of iomeprol under dynamically changing flow velocities and directions, varying hydrochemical conditions, and groundwater temperatures ranging between 12 and 17 °C.

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MATERIALS AND METHODS

Field Site and Investigation Program. The investigated stream, the Schwarzbach, is located near the city of Frankfurt, Germany, in a region that is intensely used for industrial and agricultural purposes. The aquifer beneath the stream consists of sediments with increasing grain sizes from silt at the top toward sand, and fine gravel at the bottom.11 The study site was equipped with six monitoring wells along a 2D transect following the groundwater flow direction and is directed perpendicular to the stream bank. Groundwater monitoring wells (GWMs) were installed up-gradient of the stream (GWM0) and down-gradient of the stream nearby the stream bank (GWM2a: 0.7 m distance, GWM2c: 3.4 m distance), and at a distance of 8−220 m from the stream. The piezometer network used in this work is illustrated in Supporting Information (SI) Figure SI1. Piezometric pressure heads were recorded every 10 min in the GWMs. Multilevel temperature sensors with a vertical distance of 20 cm logged the vertical groundwater temperature profiles every 10 min at a distance of 3.4 m downstream from the stream bank up to a depth of 3 m. Stream level and stream temperature were recorded every 10 min with a pressure and temperature sensor placed on the stream bed.12 Stream levels were then converted to discharge fluxes. These data types were monitored from April, 1st, 2008 until October,9th, 2010. Furthermore, during a 25-days field experiment, between September, 14th and October, 9th, 2010, every second day water samples were collected with a bufflepacker device for chemical analyses at three different depth levels (top, middle, bottom of the filter screen) within GWM0 (0.5 m, 2.5 m, 4.50 m below surface), GWM2a (0.5 m, 3.0 m, 5.50 m below surface), and GWM2c (0.5 m, 7.0 m, 14 m below surface). Additional grab samples were taken within the center of the stream. A ground-penetration radargram was collected beneath the stream bed to separate the approximately 1m-thick zone below the stream bed containing relatively freshly deposited reactive organic material such as branches or leaves, from the remaining aquifer sediments (SI Figure SI1).

rDOC;SOC, l =

⎡ ⎞ CO2 f (TC) ⎢ ⎛ ⎟ rO2⎜⎜ f (Tref ) ⎢⎣ ⎝ K O2 + CO2 ⎟⎠ ⎛ ⎞ ⎛ ⎞⎤ C NO3 K inh O2 ⎜ ⎟ ⎜ ⎟⎟⎥ + rNO3,⎜ ⎟×⎜ ⎝ KNO3 + C NO3 ⎠ ⎝ K inh O2 + CO2 ⎠⎥⎦ (1)

where rDOC and rSOC,l are reaction rates [mol L−1d−1] for DOC and SOC,l, respectively, rO2 and rNO3 are the maximum rate constants [mol L−1d−1] for aerobic and denitrifying reduction conditions that are estimated during the reactive transport model calibration, respectively. CO2 and CNO3 represent the concentrations [mol L−1] of oxygen and nitrate, respectively, 200

dx.doi.org/10.1021/es403194r | Environ. Sci. Technol. 2014, 48, 199−207


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denitrifying than under aerobic conditions,26−29 and (ii) that the biotransformation rate of iomeprol increased 2-fold when temperature was increased from 20 to 30 °C.13 For the second variant the microbially mediated transformation of iomeprol was modeled as:

KO2 and KNO3 are half-saturation concentrations, and Kinh.O2 are inhibition constants that were adopted from previous studies22 and set to 2.94 × 10−4 mol L−1, 1.55 × 10−5 mol L−1, and 1 × 10−5 mol L−1, respectively. Tref is a reference temperature that was set to 12 °C, corresponding to the annual average groundwater temperature.12 f (Tc) [°C1−] represents an empirical function for the dependence of the organic carbon mineralization on the groundwater temperature and is based on the Arrhenius equation:23 f (TC) = e

−(

1 a +a ) Tc + 273.15 1 2

reff,ICM,1.....11(T C) = (rDOC + rSOC, l + rSOC, h)kco,ICM,1...11c ICM,1...11

(4)

where reff,ICM,1...11 is the temperature-dependent modified effective first-order rate of iomeprol and its TPs [mol L −1d−1], rDOC, rSOC,l, rSOC,h are temperature-dependent degradation rates of organic matter (OM) [mol L−1d−1] kco,ICM,1...11 is the modified first-order rate constant [1/(mol L−1)] for the cometabolic transformation, and cICM,1...11 the concentration [mol L−1] of iomeprol and its TPs, respectively. First-order rate constants for variant 1 and the modified firstorder rate constant of variant 2 were estimated during the independently performed calibration of the two model variants. Due to the spatial and temporal variations of OM degradation the biotransformation reactions as implemented in variant 2 caused spatially and temporally varying half-lives. Initial and Boundary Conditions. The initial (ambient) groundwater concentrations of iomeprol, its TPs, major ions, and DOC were adopted from the concentrations that were measured toward the upstream boundary at GWM0 (SI Table SI2) at the beginning of the field experiment. For simplicity initial values were homogenously distributed throughout the model domain. The same concentrations were also used to define (i) the water composition for the SE inflow boundary, and (ii) the recharge water composition, except that for the latter it was assumed that iomeprol and its TPs were absent. SOC was set to be present throughout the model domain, whereby SOC,h was assumed to prevail within the specifically delineated zone below the streambed (SI Figure SI1). The less degradable fraction of SOC (SOC,l) was assigned to the remainder of the aquifer. The simulation period extended over 55 days (August 15th to October, 9th 2010), which includes an initial spin-up period (August 15th to September 14th, 2010). For the spin-up period only temperature data and piezometric pressure heads were available, while the time-variant streamwater composition and concentrations of iomeprol and its TPs needed to be reconstructed. For this task the mean mass fluxes of all key species including iomeprol and TPs within the Schwarzbach were computed from the 25-day long field experiment. Constant mass fluxes were assumed for the spin-up period. The mass fluxes were then converted back into daily varying concentrations by accounting for the time-varying stream discharge that was recorded for the spin-up period (SI Figure SI3). For the actual 25-day long field-experiment the measured streamwater compositions (including iomeprol and TPs) were employed in the model. During the field-experiment the hydrochemical composition of the Schwarzbach showed to mainly depend on the WWTP discharge composition. Therefore the discretely measured concentrations and computed mass loads of the ICM and its TPs are affected by a significant uncertainty. In addition further uncertainties arise from water sample collection as only grab samples were taken. Finally, the precise analytical identification of individual TPs also remains a significant challenge. To partially mitigate these uncertainties we also tracked the ratios θi of the ICM parent and TP concentration to the overall ICM concentration. The ratios θi were assumed to be less volatile within the WWTP’s mass load, and were calculated with:

(2)

where Tc is the groundwater temperature [°C], a1 and a2 are empirical constants that mimic the reaction mechanisms of the Arrhenius equation. They are adopted from literature data23 and were set to −6758.1 K and 16.1 [−]. In zones where SOC was expected to be substantially more reactive (SOC,h) the degradation rate of SOC was locally increased by a relative reactivity rreac [−]: rSOC, h = rreacrSOC, l (3) where rSOC,h is the degradation rate [mol L−1d−1] of highly biodegradable SOC,h that is estimated during the calibration and reported to reach values of up to 200.19,24 The extent of this zone within the stream bed sediments was delineated by a radargram (SI Figure SI1). Transformation of Iomeprol. Kormos et al.13 observed in batch experiments a set of 18 distinct biotransformation reactions (SI Figure SI2). They found that transformation rates were constant over time (zeroth order degradation). In contrast, Batt et al.25 reported a first-order degradation behavior. Kormos et al.13 also found that iomeprol transformation occurs in two distinct phases. During a first phase, eight TPs (TP791, TP805, TP819, TP717, TP731, TP789, TP761, and TP775) quickly emerged with TP791 forming instantaneously. After 60 days, in a second phase, seven further TPs (TP745, TP701, TP687, TP657, TP643, TP629, and TP599) were formed. Transformation reactions that showed to occur in the batch experiments13 were amended to the standard PHREEQC17 reaction database that was employed in the PHT3D simulations. However, transformation steps that involved TPs that could neither be detected in the groundwater nor in the streamwater were not considered in the reaction network. For simplification four transformation steps (k6, k7, k8, k9) were lumped from two separate reaction steps into a one-step reaction. Overall, eleven individual transformation steps were included in the reaction network (SI Figure SI2). The results of initial simulations in which a zeroth-order degradation behavior was assumed completely mismatched the field observations. Therefore the subsequent modeling focused on two additional variants. The first variant assumed standard first-order kinetics for all transformation reactions, thereby ignoring the influence of the dynamically changing geochemical conditions and temperatures on the reaction rates. The second tested variant assumed that first-order degradation rates could vary in space and time, depending on local geochemical conditions and groundwater temperatures. The latter, more complex model formulation was used to explore whether the bulk of the iomeprol transformations may occur as a temperature-dependent cometabolic process. This hypothesis was based on previously reported laboratory observations that detected that (i) transformation rates were 10 times higher for 201



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Figure 1. Measured (symbols) versus simulated (lines) oxygen and nitrate concentrations along the stream bank. Nonreactive transport simulations (dashed lines) are compared with reactive transport simulations (green and blue lines). Computed degradation rate of organic matter (DOC, SOM,l, SOM,h) by nitrate reduction (black line). Losing and gaining conditions are derived from the gradient between the stream stage and GWM2a (SI Figure SI1)12.

θi =

ci c parent + ∑n = 1,max TP c TP, n

trations that were measured at GWM2a and GWM2c. Measured ratios θ for iomeprol were summarized into TPs of phase I and II (SI Figure SI2) and trialed as alternative calibration constraints to concentration measurements. Long-Term Fate of Iomeprol. Based on the calibrated model additional simulations were conducted to assess the long-term fate and transport characteristics of iomeprol and its TPs near the stream/groundwater-interface. For the underlying flow simulations we employed transient, measured daily data for the stream stage, piezometric pressure heads at the SE inflow and NW outflow boundary, from the period between April, 1st 2008 and October 9th, 2010. The initial geochemical conditions were assumed similar to those employed to simulate the monitored study period. The time-varying streamwater composition was reconstructed as previously described for the spin-up period. The long-term simulations compare the ICM mass loads and plumes as obtained by comparative nonreactive and reactive transport simulations for both reactive model variants.

(5)

where θi is the ratio of the individual ICM (parent or TP) to the overall ICM concentration, cparent is the concentration of the ICM parent, cTP is the concentration of all TPs, n is the number of TPs, and ci is the concentration of the individual ICM. Model Calibration. The calibration of the conservative model11 provided a robust description of the flow, solute and temperature transport processes and reliable estimates of hydraulic and conservative transport model parameters. For the present study the set of estimated physical model parameters (hydraulic conductivity, porosity, dispersion, thermal conductivity, specific heat) was not further modified. In the reactive transport simulations kinetic parameters (rO2, rNO3, rreac.) controlling DOC/SOC mineralization were estimated and constrained by DOC, DO, and nitrate concentrations measured at GWM2a and GWM2c. It appears that neither reduction of iron nor manganese occurred, which is supported by the fact that the conservative model could reproduce the measured iron and manganese concentrations and also due to the continuous presence of at least residual concentrations of nitrate. The calibration of the reactive transport model allowed estimating first-order rate constants for iomeprol and its TPs for both variants for every transformation reaction. The estimates were constrained by iomeprol and TPs concen-

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RESULTS AND DISCUSSIONS Redox Zonation. After model calibration, reactive transport simulations in which organic carbon degradation was the main driver for spatial and temporal changes in redox zonation. The simulations could reproduce the level of the measured DO and nitrate concentrations at most locations of the field site reasonable well. Due to the gaps in surface water concentration measurements during the spin-up period it was difficult to 202



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Figure 2. Measured (circles) and simulated (lines) ratios and concentrations of iomeprol and transformation products close to the stream bank (top level of GWM2a) versus reactive transport model results (solid red line) and nonreactive transport simulation results (dashed red line). Iomeprol and TPs concentrations and ratios measured in the stream (green triangles) and simulated groundwater temperatures at GWM2a (blue line).

Table 1. Reactive Transport Model Estimates: Calibrated Values for the Degradation of Organic Carbon, Calculated First-Order Rate Constant for the Degradation of Iomeprol (Variant 1) And Variable Half-Lives Using a Modified Temperature-Dependent First-Order Co-Metabolic Degradation of Iomeprol (Variant 2) Versus Half-Lives Derived from Aerobic Batch Experiments13 calibrated values for the degradation of organic carbon parameter

DOC

DT f

g

calibrated values mean /min iomeprol parent into TP 791 TP791 into TPs phase I TPs of phase I into phase II

SOC,l

3.90 × 10−5 [mol L−1 d−1] 1.20 × 10−6 [mol L−1 d−1] 8.0 [−] biodegradation of iomeprol

rO2 rNO3 rreac

50 h

7.80 × 10−5 [mol Lbulk−1 d−1] 1.98 × 10−5 [mol Lbulk−1 d−1]

k [mol/days]

[days] aerobic batch experiments

108.4/68.7 108.0/68.6 71.3/20.7

(eb)

(jb)

6.3 /21.3 15.4(ec)/42.4(jc) 20.6(ed)/20.6(jd)

a

calibrated values −7

5.0 × 10 1.0 × 10−4 8.0 × 10−4

i

aerobic batch experimentsa 2.0 × 10−6b 7.8 × 10−7c 6.1 × 10−7d

Transformation of the iomeprol parent into TPs of phase II as measured in Kormos et al. (2010). bSoil incubated with wastewater, 30 °C. cStream bed sediment (high organic carbon content of 2.3%). dSoil without prior irrigation, 20 °C. eDT50 linearfor zero-order degradation. fEstimated with the reactive transport model assuming a first-order biodegradation (variant 1). gMean half-life using DT50 = ln(2)/keff, ICM 1...11. hSmallest half-life calculated during the high water event with maximum nitrate infiltration and groundwater temperatures of 17.5 °C resulting in maximum organic carbon degradation rates. iEstimated with the reactive transport model kinetically assuming a temperature-dependent first-order co-metabolic biodegradation (variant 2). jDT50 totalthat Value includes the lag phase. a

match the detailed temporal evolution of the observed redoxspecies concentrations (Figure 1). Sensitivity runs showed that the impact of the estimated initial concentrations at the start of

the spin-up period on calculated oxygen and nitrate concentrations during the study period remained small. Most of the DO and nitrate contained in the streamwater that 203



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an average of 0.07. Highest ratios were measured for TP687, TP643, TP629, and TP599 (θ = 0.30, 0.37, 0.62, and 0.4), respectively. The maximum concentration were detected for TP 629 with 1.55 μg L−1 and TP599 with 1.94 μg L−1, whereas only one of them (TP629) was found at an elevated ratio and concentration (1.0 μg L−1) in the Schwarzbach stream. This suggests that TP687, TP643, and TP599 were produced during transport and residence within the hyporheic zone. At GWM2c the highest detected θ for iomeprol was 0.14 while TP687, TP643, TP629, and TP599 were detected at ratios of 0.28, 0.27, 0.43, and 0.40, respectively. Also, TPs of phase II prevailed in the riparian zone at ratios that were above their initial ratios within the streamwater. This suggests that biotransformation of iomeprol also proceeds under the predominantly anaerobic, denitrifying conditions, that prevailed at the surface water/ groundwater-interface. It can be seen that for iomeprol and TPs of phase II major features simulated and measured breakthrough curves agree in terms of their ICM ratios, while it was difficult to match ICM concentrations. This difficulty is mainly attributed to the uncertainty in the WWTPs ICM mass load into the streamwater. However, Figure 2 demonstrates the distinct differences that are obtained for iomeprol between reactive and nonreactive transport simulations. For the TPs of phase I, however, the reactive simulations do not provide a distinctly better match to neither the measured ratios nor concentrations (illustrated for TP 805) than the nonreactive simulations. This could indicate that under the investigated field-scale conditions a direct transformation step from iomeprol into TPs of phase II has occurred. In contrast, as also shown in Figure 2 for the top part of GWM2a, the reactive transport simulations for the TPs of phase II approximate the measured concentrations closer than the nonreactive simulations. More convincingly though, if the simulated and measured θs of the phase II TPs are compared instead of concentrations, the reactive transport simulations provide a clearly better match then the conservative simulations. This is illustrated in Figure 2 for TP629, which was produced at the highest rates among all TPs of phase II. The transformation rates for iomeprol into TPs of phase I as employed in the model for variant 2 under the predominantly anaerobic conditions correspond (at 12 °C) to a half-life of ∼108 days. At laboratory scale13 aerobic half-lives rangedlag phases included between 20.6 and 42.4 days (Table 1) and after 60 days 15 transformation products were identified. Modeled biotransformation of TPs of phase I into phase II resulted in a minimum half-life of ∼21 days, which is in the same range as the laboratory-scale half-life measured at 20 °C for untreated soils under aerobic conditions. Anaerobic rates are therefore distinctively below those determined under aerobic conditions. Impact of Model Complexity and Parameter Sensitivities. The above-discussed results were derived with the reaction model which assumed co-metabolic ICM degradation coupled to the temperature-dependent organic carbon mineralization rate (variant 2). The additional model complexity, as compared to the simpler variant 1, reduced the mean residuum for iomeprol by 32% and for TPs of phase II by 67%. However, a closer inspection of the geochemical variability during the 25-days field experiment shows that denitrifying conditions were dominating in the zone where most ICM transformations took place. This explains that under the prevailing conditions of relatively constant denitrification rates

Figure 3. Simulation of long-term trends (April, 1st, 2008 to October, 9th, 2010) of iomeprol ratios at GWM2c (3.4 m from the stream bank) that compare reactive transport model results using variant 2 (solid read line), variant 2 at isothermal conditions of 14 °C (dasheddotted blue line), and variant 1 (dashed green line) with nonreactive transport models results (dashed-dotted orange line) versus measured iomeprol ratios (circles) obtained during the field experiment (September 14th, 2010 to October, 9th, 2010). Top plot gives measured stream temperature (blue line) and computed infiltration velocity at shallow parts of GWM2c (magenta line, negative values indicate stream-gaining conditions).

entered the groundwater across the streambed was relatively rapidly consumed during degradation of reactive SOC,h. Both, measured and simulated DO concentrations were below 4 mg L−1 close to the stream bank (GWM2a) and below 2 mg L−1 at all times at the monitoring locations situated more than 1 m downstream from the stream bank (GWM2c). In contrast, in the corresponding nonreactive transport simulations DO and nitrate concentrations were not significantly attenuated by mixing/dilution and remained clearly overestimated. Fate of Iomeprol at the Surface Water-Groundwater Interface. During the study period iomeprol was detected at concentrations of up to 1.2 × 10−9 mol L−1 or 0.93 μg L−1 (θ = 0.3) in the Schwarzbach stream. Two TPs (TP629, TP745) briefly reached similar concentrations of up to 1.1 × 10−9 (θ = 0.36) and 1.3 × 10−9 mol L−1 (θ = 0.41), respectively (SI Figure SI3). Seven further TPs (TP805, TP791, TP717, TP701, TP687, TP643, TP599) were detected in the streamwater at concentrations of up to 5 × 10−10 mol L−1 (θ = 0.2). The presence of these TPs in the stream suggests that they have already formed during the wastewater treatment process. Five TPs (TP819, TP789, TP775, TP761, TP731, TP657) that were identified in the batch experiments of Kormos et al.13 were not detected in the stream. Figure 2 shows the comparison between simulated and measured iomeprol and TP concentrations for GWM2a, together with the respective ratios θ. In the groundwater at GWM2a the maximum iomeprol concentrations reached 0.47 μg L−1, and θ ranged for iomeprol between zero and 0.28 with 204



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Figure 4. Simulated long-term mass flux at the surface water/groundwater interface with a reactive transport model using variant 2 (red line for iomeprol and blue line for TP629) versus nonreactive transport model results (orange line for iomeprol and green line for TP629). Simulated plumes of iomeprol and TP 629 after (A) alternating conditions (left figures), and (B) an intense flooding event (right figures).

(Figure 1) and only small temperature variations (△T ≤ 5 °C) simulated ICM transformations during the calibration period varied only modestly among the model variants. The model calibration showed parameter sensitivities to be highly variable. Selected values for KO2, KNO3, and Kinh.O2 had a minor impact on the model results and were thus set to literature values.22 Moreover, due to the narrow temperature range during the field experiment the sensitivity of a1 and a2 was low. Kinetic parameter rO2 and rNO3 that controlled the degradation of DOC and SOC,l were moderately sensitive and were found to be in the same range as reported literature values (Table 1). High sensitivities were observed for rreac that drives the degradation of SOC,h and limits the inflow of nitrate into the groundwater. First-order rate constants for the biodegradation from TP791 into TPs of phase I and into phase II showed a considerable impact on the simulation results. However, the highest sensitivity and most relevant parameter for the model calibration was the first-order rate constant for the transformation from iomeprol to TP791. Long-Term Fate of Iomeprol. To assess the longer-term transport characteristics of iomeprol under the highly transient flow conditions the entire hydraulically monitoring period was simulated with the reactive transport model. The simulations started with an extended period of stream-losing conditions, followed by a period of generally stream-gaining conditions which were only occasionally interrupted by short losing periods. After 500 days, a long-term stream-losing period developed, which continued over 300 days. From 800 days until the end of the simulation period, the stream-gaining conditions were again interrupted by short losing periods. Days 895 to 920 of the long-term simulations coincide with the 25-day long intense monitoring period (field experiment).

Figure 3 shows the simulated breakthrough curves for iomeprol at GWMC2c, that is, at 3.4 m distance from the stream bank. As can be seen, clear differences between nonreactive and reactive transport simulations emerge specifically during periods of losing conditions. For the long-term simulations the discrepancy between the two variants of the reaction models becomes much more distinct. The main reason for this discrepancy is the higher variability of the nitrate inflow over the year driving the denitrification rate and higher fluctuations of the streamwater temperatures, which affect the iomeprol transformations much stronger than during the relatively short model calibration period. Comparing temperature-dependent and isothermal co-metabolic biotransformation identifies that during hot summer periods, when the streamwater temperature reaches 25 °C, biodegradation is slightly enhanced (Δθ = 0.02). During cold winter periods the low stream temperatures of 5 °C result in a distinctive limitation of the co-metabolic biodegradation (Δθ = 0.18). Figure 4 illustrates simulated concentration contours for iomeprol and TP629 after an extended period of alternating conditions (A) that was overall dominated by stream-gaining conditions (t = 400 days), and after long-term stream-losing conditions (t = 800 days) (B). Both iomeprol and TP629 concentrations decrease to LOQ after long-term alternating conditions. In contrast, after long-term stream-losing conditions TP629 is produced at sufficiently high rates and occurs then several meters downstream with the regional groundwater flow, while iomeprol persists only in proximity of the stream bank. Thus, during the stream-losing period, the maximum TP629 mass flux increases to 3 × 10−8 mol d−1, while the average mass flux of iomeprol is 10-fold smaller. 205



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The difficulty to match measured iomeprol and TPs concentrations/ratios is mainly attributed to the uncertainty of the ICM fluctuations in the streamwater during the spin-up period. Future investigations of ICM reactive transport at the stream/groundwater interface should therefore sample iomeprol within the surface water more often and for a longer period, while collecting data at a high frequency in deeper multilevel wells may be less useful. Both data and model results highlight the challenges in identifying and quantifying the ICM transformation reactions under the highly variable hydraulic, temperature and geochemical conditions. While not underpinned by measured, longer time-series, it is evident that the long-term behavior of the compounds is more accurately quantified with a model of higher complexity.

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ASSOCIATED CONTENT

S Supporting Information *

Additional information as noted in the text. This material is available free of charge via the Internet at http://pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Author

*(I.E.) E-mail: [email protected]. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS Thanks are due to Christoph Kludt (TU Darmstadt, Germany) for collecting the field data. We thank Uwe Kunkel (Federal Institute of Hydrology, Germany) for his numerous and very helpful comments. The laboratory analysis of the ICM was conducted at the Federal Institute of Hydrology, Germany, and financially funded by the German Federal Ministry of Environment, Nature Conservation, and Nuclear Safety (BMU).

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REFERENCES

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Reactive Transport of Iomeprol during Stream-Groundwater Interactions

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