Interactions and Reductive Reactivity in Ternary Mixtures of Fe(II

May 29, 2019 - 21. Elsner, M.; Schwarzenbach, R. P.; Haderlein, S. B. Reactivity of Fe(II)-bearing minerals toward reductive transformation of organic...
0 downloads 0 Views 854KB Size
Download PDF
Article Cite This: Langmuir 2019, 35, 8220−8227

pubs.acs.org/Langmuir

Interactions and Reductive Reactivity in Ternary Mixtures of Fe(II), Goethite, and Phthalic Acid Based on a Combined Experimental and Modeling Approach Jianzhi Huang,† Qihuang Wang,‡ Zimeng Wang,‡,§ and Huichun Zhang*,† †

Department of Civil Engineering, Case Western Reserve University, Cleveland, Ohio 44106-7220, United States Department of Environmental Science and Engineering, Fudan University, Shanghai 200086, China § Shanghai Institute of Pollution Control and Ecological Security, Shanghai 200092, China Downloaded via KEAN UNIV on July 21, 2019 at 03:11:13 (UTC). See https://pubs.acs.org/sharingguidelines for options on how to legitimately share published articles.

‡

S Supporting Information *

ABSTRACT: The interactions between organic ligands, Fe(II), and iron oxides are important in biogeochemical redox processes. The effect of phthalic acid (PHA) on the reductive reactivity of Fe(II) associated with goethite was examined using batch adsorption and kinetic studies, attenuated total reflectance−Fourier transform infrared spectroscopy (ATR−FTIR), and surface complexation modeling (SCM). PHA significantly inhibited the reductive reactivity of Fe(II)/goethite, as quantified by the pseudo-first-order reduction rate constants (k) of p-cyanonitrobenzene. The k value decreased from 1.68 ± 0.03 to 0.338 ± 0.14 h−1 at pH 6.0 as the PHA concentration increased from 0 to 1000 μM. The effects of the co-adsorption of Fe(II) and PHA onto goethite were then investigated to study the inhibition mechanism. The adsorption experiments showed that Fe(II) slightly enhanced PHA adsorption, whereas PHA did not affect Fe(II) adsorption, suggesting that the inhibition was not due to different amounts of Fe(II) adsorbed. The ATR−FTIR spectra of the adsorbed PHA in the ternary mixtures demonstrated that the major surface species was outer-sphere species, with minor inner-sphere complexes formed. SCM results showed that the presence of PHA (L) led to the formation of a type A ternary species ((FeOFe+)2···L2−) on the goethite surface, decreasing the abundance of the reactive species (FeOFeOH). Moreover, the adsorption of PHA on the surface of goethite might block the reactive sites and inhibit the electron transfer between Fe(II) and goethite, thus decreasing the reactivity. Overall, these findings provided new insights into the reaction mechanisms of surface-adsorbed Fe(II), which will facilitate the development of new technologies for site remediation and more accurate risk assessment.

■

INTRODUCTION In anoxic environments, the Fe(II)/Fe(III) redox reaction plays an important role in biogeochemical redox processes, for example, the transformation of pollutants in the environment.1−8 Over the past 3 decades, Fe(II) complexed with iron oxides (or surface-bound Fe(II)), an environmentally important reductant, has been extensively investigated for its ability to reduce a large number of contaminants.9−12 Fe(II) complexation with iron oxides can significantly enhance the reductive reactivity by lowering the redox potential of Fe(II).1,13 Numerous factors can affect the reductive reactivity of the surface-bound Fe(II), including pH,9,10,14−16 Fe(II) concentration,9,16 Fe(II) surface speciation,11−13,17,18 second metal oxides,19 and properties of the mineral oxides.10,20−22 A common observation is that the reduction kinetics strongly depended on the extent of Fe(II) adsorption, which was largely influenced by both solution pH and the soluble Fe(II) concentration.23 Despite the above significant body of work, little is known about how ligands affect the reductive reactivity of Fe(II)/iron oxides. Previous studies explored the reductive reactivity of Fe(II) with iron oxides and the electron transfer between them in the © 2019 American Chemical Society

presence of inorganic ligands and natural organic matter (NOM). One study showed that phosphate, silicate, carbonate, and NOM did not inhibit the electron transfer and phosphate did not have any effect on the atom exchange between aqueous Fe(II) and the Fe(III) in goethite (but the sorption of a longchain phospholipid completely shut down the electron transfer).24 Another work, on the contrary, reported decreased extents of atom exchange between aqueous Fe(II) and an Fe(III) mineral in the presence of silicate and explained it based on the adsorption of Si which inhibited direct adsorption of Fe(II) onto the mineral surface. 25 Hinkle et al. demonstrated that phosphate and sulfate can increase the extent of Fe(II) adsorption onto iron oxides.26 However, they speculated two conflicting hypothesis: (1) the reductive reactivity of Fe(II)/iron oxides was promoted due to the increasing amount of Fe(II) adsorbed; (2) ligands might stabilize Fe(II) on the surface to inhibit electron transfer and Received: February 22, 2019 Revised: April 26, 2019 Published: May 29, 2019 8220

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227

Article

Langmuir

attenuated total reflectance−Fourier transform infrared spectroscopy (ATR−FTIR) and surface complexation modeling (SCM). This work offered new insights into the role of carboxylic acids in environmental transformation processes. The findings will improve our understanding of the effects of LMW carboxylic acids on the reductive reactivity of Fe(II)/ iron oxides, which will be useful in the development of predictive models for the fate and transformation of contaminants.

atom exchange. Therefore, further evidence is required to examine which hypothesis is correct. In addition to inorganic ligands, NOM has been reported to inhibit the reductive reactivity of Fe(II)/iron oxides due to (1) NOM adsorption to block the access of Fe(II) to the surface reactive sites and (2) NOM complexing with or oxidizing the surface Fe(II).27,28 The reductive reactivity was well correlated with eight NOM physiochemical properties, including molecular weight, carboxyl concentration, and carbon, oxygen, nitrogen, aliphatic, heteroaliphatic, and aromatic content.27 Despite the above studies, there is still a lack of mechanistic understanding of what types of surface species have formed in the presence of ligands and how ligands affect the reductive reactivity of Fe(II)/Fe oxides. A number of studies have examined the co-adsorption of metal ions and ligands onto metal oxides. It has been shown that ligands may influence the adsorption of metal ions onto metal oxides29−34 and that metal ions could also affect the adsorption of ligands.35,36 However, the effects of ligands on the adsorption of metal ions cannot be readily generalized, as ligands in some cases promoted metal ion sorption,33,37,38 whereas in other cases inhibited metal ion sorption.39,40 This indicates that the effects of ligands on metal ion adsorption are system-dependent and vary with pH and the nature and ratio of the metal ions and the ligands involved.41 Among the various ligands reported, low-molecular-weight (LMW) carboxylates, originated from both natural and anthropogenic sources, form an important class of ligands in aquifers and soils.42,43 For example, the concentration of phthalic acid (benzene-1,2-dicarboxylic acid, PHA) was 2−880 μM in landfill leachates.44 Because of their ability to form complexes with metal ions both in solution and on mineral surfaces, carboxylates can influence a number of geochemical processes,43 such as the dissolution and weathering of mineral oxides45−47 and the fate and transport of contaminants in the environment.48,49 For example, PHA is known to interact with both metal ions and metal (hydr)oxides,30,50,51 which can influence the sorption of metal ions on iron oxide surfaces through different mechanisms.52 However, neither coadsorption of Fe(II) and LMW carboxylic acids nor the effect of LMW carboxylic acids on the reduction of contaminants by Fe(II)/iron oxides has been studied. In addition, LMW carboxylic acids contain −COOH, which is an important functional group in NOM. Because PHA is an important LMW carboxylic acid,51,53,54 developing insight into Fe(II)-PHA speciation on iron oxide surfaces and how it affects the redox reactivity would help understand the effects of NOM on similar reductive reactions. The aim of this study was to explore the speciation in ternary PHA, Fe(II), and goethite mixtures and to evaluate how the species affected the reductive reactivity of Fe(II)/ goethite. To achieve this goal, p-cyanonitrobenzene (pCNB), a commonly used probe compound for reduction because of its facile reactivity and poor adsorption onto mineral surfaces,55,56 was examined for its reductive kinetics as an indicator for the Fe(II)/goethite reactivity under different conditions. To examine how PHA affected the Fe(II)/goethite reactivity, we conducted sedimentation experiments for the extent of aggregation under pseudo-steady state conditions (higher extents of aggregation were related to lower reactivity57), examined the change in the amount of Fe(II) adsorbed under different pHs and PHA concentrations, and studied the types of surface Fe(II) species formed on goethite surfaces using

■

EXPERIMENTAL METHODS

Reagents. All chemicals were used as received. pCNB was purchased from Aldrich. Other chemicals, including acetonitrile, FeCl2·4H2O, ferrozine, HCl (trace metal-grade concentrated), 4morpholinepropanesulfonic acid, 2-(N-morpholino)ethanesulfonic acid (MES), NaOH, sodium acetate, and PHA, were obtained from Fisher or Sigma-Aldrich. Goethite (Bayferrox 910) was obtained from Lanxess. Analytical Methods. The concentrations of PHA in the supernatant were analyzed by an Agilent 1200 reversed-phase highperformance liquid chromatography (HPLC) system with a diode array detector and a Zorbax XDB-C18 column (4.6 × 150 mm, 5 μm) at a flow rate of 1 mL/min with dilute phosphoric acid at pH 3.00 and methanol (60:40). For pCNB, the same mobile phase was used but gradient elution was run to separate pCNB and its products. Aqueous Fe(II) and total Fe were analyzed by a modified ferrozine method.4,49,58 Briefly, 100 μL of reagent A (10 mM ferrozine with 0.1 M ammonia acetate) was added to 1 mL samples. The absorbance was measured by a UV−visible spectrophotometer (Agilent 8670) at 562 nm for Fe(II) concentrations. Reactor Setup. DI water used to prepare all solutions was degassed by boiling under vacuum before being transferred into an anaerobic chamber with an atmosphere of 95−98% N2 and 2−5% H2 (Coy Laboratory Products, Inc.). All experiments were conducted in 50 mL amber glass serum bottles under constant stirring. NaCl (0.1 M) was added to maintain ionic strength. Reactors were prepared by adding goethite to DI water-containing buffer (20 mM MES buffer for pH 5.5−6.5 and 20 mM MOPs buffer for pH 7.0−8.0) and NaCl. The reactors were then equilibrated overnight to ensure sufficient hydration of goethite surfaces. After that, certain amounts of ligands and Fe(II) were added (reaction conditions at pH 6: NaCl 0.1 M, 20 mM MES, goethite: 5 g/L, Fe(II): 250 μM, PHA: 0−1000 μM and reaction conditions at pH 7: NaCl 0.1 M, goethite: 0.2 g/L, Fe(II): 100 μM, PHA: 0−1000 μM), which were then equilibrated for 24 h at 25 ± 2 °C. Reactions were initiated by adding a certain amount of pCNB to the reactors. Aliquots of suspensions were collected and filtered using 0.2 μm nylon filters to quench the reaction. Then, aqueous pCNB was analyzed by HPLC. Rate constants (k) of the reductive reactivity were obtained based on the pseudo first-order kinetics: ln(C/C0) = −kt, where C0 is the initial pCNB concentration and C is the pCNB concentration at time (t). ATR−FTIR Experiments. ATR−FTIR spectra were collected with a PerkinElmer Spectrum 100 FTIR spectrometer, which was equipped with a deuterated triglycine sulfate detector. A horizontal attenuated total reflectance attachment and a 45° ZnSe crystal mounted in a flow cell (PIKE Technologies, USA) were used. For each sample, 500 scans with a spectral resolution 4 cm−1 were taken. Spectrum Software (Version 10.4.2, PerkinElmer Inc.) was used for data collection, smoothing, and spectral calculation. To prepare samples for the ATR−FTIR analysis, 1 mL of 1 g/L goethite suspension was spread on the surface of the crystal, after which the crystal was dried in a vacuum oven at 40 °C for 1 h.59 Following that, the crystal was gently rinsed with deoxygenated DI water for several times in order to remove loosely deposited particles. NaCl solution (0.1 M) at a predetermined pH was passed through the flow cell at a rate of 0.3 mL/min for at least 5 h. The background spectrum was then collected that consisted of the absorbance of the ZnSe crystal and the deposited goethite. The solution was then 8221

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227

Article

Langmuir Table 1. Equilibrium Reactions and Constants for Surface Species (L = PHA)a ψ0

equilibrium reactions

≡FeOH + H+ ⇋ ≡ FeOH 2+

≡FeOH ⇋ ≡ FeO− + H+

(2) (3)

≡FeOH + Fe2 + ⇋ ≡ FeOFe+ + H+

(4)

≡FeOH + Fe2 + + H 2O ⇋ ≡ FeOFeOH + 2H+

(5)

2(≡FeOH) + L2 − + 2H+ ⇋ ≡ (FeOH 2+)2 ...L2 − 2−

+

2(≡FeOH) + L

+ 2H ⇋ ≡ Fe2L + 2H 2O

2(≡FeOH) + 2Fe

2+

2−

+L

+

(6) (7)

2−

⇋ ( ≡ FeOFe )2 ...L

+ 2H

+

(8)

ψβ

log K

refs

+1

0

6.7

65

−1

0

−9.0

65

+1

0

0

0

−9.2

0.27

this study

this study

+2

−2

21.0

this study

0

0

16.8

this study

+2

−2

5.6

this study

Note that ψ0 and ψβ denote the Coulombic term for the 0 and β layer, respectively.

a

Figure 1. k values under different PHA concentrations (0−1000 μM) at pH 6 (a) and pH 7 (b). Reaction conditions: (a) for pH 6, NaCl 0.1 M, 20 mM MES, goethite: 5 g/L, Fe(II): 250 μM; (b) for pH 7, NaCl 0.1 M, 20 mM MOPs, goethite: 0.2 g/L, Fe(II): 100 μM. changed to 50, 100, 200, 500, or 1000 μM Fe(II) solution in the presence of 0.5 mM PHA at the same pH. All the above preparation was done in the anaerobic chamber. Macroscopic Fe(II) and PHA Adsorption. A series of adsorption experiments were conducted to investigate the co-adsorption of Fe(II) and PHA. The experimental conditions were similar to those of the kinetic experiments. Solution pH ranged from 5.5 to 8.0 with intervals of ∼0.5. MES buffer was used for pH 5.5−6.5, while MOPs buffer for pH 7−8. Certain amounts of PHA and Fe(II) were added and equilibrated for 24 h. After that, the suspensions were filtered through 0.2 μm nylon filters, and the concentrations of Fe(II) and PHA in the supernatants were analyzed by the ferrozine method and HPLC. Sedimentation. Sedimentation kinetics of the goethite with or without PHA at pH 6 were monitored by measuring the absorbance at 508 nm using a UV−vis spectrophotometer (Agilent 8453) at different times.49 Here, it was to measure the scattering of light rather than absorption of light. Normalized absorbance measurements (A/ A0) were reported as a function of time. The mixtures of goethite and PHA were stirred on a magnetic stir plate and equilibrated for 24 h before the experiments, because this was the pre-equilibrium time used for the kinetic experiments. The purpose was to determine if PHA affected the extent of aggregation of the goethite particles. Surface Complexation Modeling. Data from the adsorption experiments were modeled based on a triple-layer model using MINEQL+5.0.60,61 The surface area of goethite was measured to be 15 m2/g, which agrees with the published data. The parameters of the commercial goethite used (Bayferrox 910) for the modeling were obtained from previous research62 and are listed in Table S1. The capacitance of the inner charge-free layer (C1) and that of the outer charge-free layer (C2) were optimized based on eq 1 1 CStern

=

1 1 + C1 C2

The adsorption of Fe(II) and PHA was separately modeled to identify the surface reactions and equilibrium constants needed in the model. Two surface species (FeOFe+ and FeOFeOH) that have been used in previous studies were employed to model the Fe(II) adsorption (eqs 4 and 5 in Table 1).11,17 For PHA adsorption, the surface complexes were selected based on the prior modeling and IR spectroscopy studies,51,63 which suggested a dominant outer-sphere species with minor inner-sphere complexes on iron oxides (eqs 6 and 7). After constraining the equilibrium constants for reactions 4−7 for Fe(II) or PHA as the single adsorbate, these binary surface complexes were applied to the co-adsorption experiments. One ternary surface complexation reaction was then incorporated in the model to improve the fitting (eq 8). With the equilibrium constants for the binary surface complexation reactions maintained constant at this stage, the method of least-square residuals was used to obtain an optimum equilibrium constant for the ternary surface complex.26,64

■

RESULTS AND DISCUSSION

Effects of PHA Concentration on pCNB Reductive Transformation. A set of experiments was performed with various PHA concentrations to evaluate the influence of PHA concentration on the kinetics of pCNB degradation. The rate constant values (k) of pCNB reduction decreased from 1.68 ± 0.03 to 0.34 ± 0.14 h−1 when the PHA concentration increased from 0 to 1000 μM at pH 6 (Figures 1a and S1). This is similar to previous studies when adding NOM to Fe(II)/goethite.27,28 No significant pCNB reduction was observed in Fe(II) alone controls (data not shown), similar to previous studies.9,10 In addition, control batch reactors were prepared and no significant pCNB reduction was observed in solutions without goethite, indicating that PHA or Fe(II)-PHA complexes did not reduce pCNB over the time of interest. The effect of PHA on the reduction reactivity of Fe(II)/ goethite at pH 7 was similar to that at pH 6 (Figure 1b). The k

(1)

The involved surface reactions are listed in Table 1. 8222

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227

Article

Langmuir values decreased from 2.45 ± 0.03 to 1.37 ± 0.57 h−1 when adding PHA from 0 to 1000 μM. The inhibition by PHA was more obvious at pH 6 than at pH 7. This result was most likely due to the difference in the amounts of PHA adsorbed at different pHs (Figure S2). The adsorption of anionic PHA by metal oxides at higher pHs has been reported to be negligible because of electrostatic repulsion between the anions and the increasingly negatively charged surfaces.51,52 In addition, numerous studies have shown that the reduction kinetics strongly depended on the amount of Fe(II) adsorbed.23 However, the amount of Fe(II) adsorbed was not much affected by the addition of PHA (Figure S2); thus, the decrease in the reactivity was not because of change in the amount of Fe(II) adsorbed. PHA here might have affected goethite aggregation state and/or the adsorbed Fe(II) species, as explained below. Sedimentation. The aggregation state of metal oxide has been reported to affect the reductive reactivity.23,66 Therefore, sedimentation experiments were conducted to see the extent of aggregation (note, not to measure aggregation kinetics) between oxide particles after adding different concentrations of PHA. Unlike what we observed earlier regarding the effect of Aldrich humic acid on the aggregation of goethite,48,49 PHA did not have any significant effect on the sedimentation state of goethite (Figure S3). This result rules out the possibility of PHA to inhibit the reductive reactivity by changing the goethite aggregation state. In fact, it has been shown that the inhibition of Fe(II)/goethite reactivity by NOM was not due to the difference in the aggregation states.27 Based on the above results, the addition of PHA might have affected the adsorbed Fe(II) species to result in the decrease in the reductive reactivity. In order to understand what species formed on the goethite surface, ATR−FTIR and SCM were examined, as follows. ATR−FTIR Spectroscopy. PHA has two pKa values (pKa,1 = 2.89, pKa,2 = 5.41); thus, it may exist in three different protonation states as a function of solution pH, including a fully deprotonated species (L2−), a singly protonated species (HL−), and a neutral species (H2L) (Figure S4). The ATR− FTIR spectra of aqueous PHA were in good agreement with those previously reported51 and exhibited systematic differences as the relative amounts of the three species shifted with pH (Figure 2a). At low pH, the peaks at 1290 and 1710 cm−1 were due to the stretching vibrations νC−OH and νCO, respectively.51 As pH increased, three major peaks at 1553, 1403, and 1384 cm−1 appeared. The peak at 1553 cm−1 was assigned to the asymmetric stretching vibration (νas) of the deprotonated carboxylate groups (COO−), while the peaks at 1403 and 1384 cm−1 were assigned to the symmetric stretching vibration (νs) of the deprotonated carboxylate groups (COO−).51 These features increased in intensity as the pH increased, and at the higher pHs studied (pH ≥ 6), they dominated the spectrum of the solution composed of 89.4% L2− and 10.6% HL− (Figure 2a). The adsorption of PHA on metal (hydr)oxides has been extensively studied, and the mode of adsorption has been described as inner-sphere and/or outer-sphere. For example, a fully deprotonated, outer-sphere complex was typically the major species, whereas two additional fully deprotonated inner-sphere complexes were observed at low pH when PHA adsorbed on hematite surfaces.51 In the binary system of PHA−goethite (Figure 2b), the bands at 1710 cm−1 (νCO) and 1290 cm−1 (νC−OH) were not evident in the system,

Figure 2. (a) ATR−FTIR spectra of aqueous PHA (20 mM) as a function of solution of pH; (b) spectra of PHA (0.5 mM) adsorbed onto goethite at pH 6 in the presence of Fe(II) (0−1000 μM) in 0.1 M NaCl solution.

indicating that the adsorbed PHA was in the fully deprotonated form.51,67 The spectra of the binary system (PHA−goethite) were roughly similar to those of aqueous phthalate, suggesting a dominant outer-sphere mode of adsorption. However, there are some small differences in the spectra, indicating that there were likely inner-sphere species formed on the goethite surface, as explained below. First, the peak at 1553 cm−1 appears to shift slightly to a higher wavenumber (1563 cm−1) at pH 6 (Figure 2b). Second, the peak at 1384 cm−1 disappeared and the peak at 1403 cm−1 moved to a higher wavenumber. All these changes could be ascribed to the presence of inner-sphere species formed on the goethite surface.51 It should be noted that the asymmetric COO− stretching region is about 1660−1450 cm−1, which might be affected by water peaks.67 Therefore, the symmetric COO− stretching region (1460−1300 cm−1) was used for the accuracy of analysis. In the ternary systems of Fe(II)−PHA− goethite, the symmetric peaks with increasing Fe(II) concentration appeared to be nearly the same as those without Fe(II), indicating that Fe(II) did not change the coordination mode of PHA. Therefore, the increase in the amount of PHA adsorbed with increasing Fe(II) might be due to the ternary outer-sphere adsorption through electrostatic interactions. In addition, it should be mentioned that based on previous research, inner-sphere species can only form at lower pH.63 Therefore, in the ternary systems at pH 6, the major species of the complex adsorbed on the goethite surface was outer-sphere species, with minor binary inner-sphere complexes formed. Macroscopic Sorption. Before using SCM to examine the adsorbed Fe(II) species, the adsorption of Fe(II) and PHA at different pHs was obtained first. Although recent studies have 8223

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227

Article

Langmuir

As can be seen, the adsorption of PHA was also pH-dependent. Maximum adsorption occurred at pH 5.5, while there was little adsorption at basic pH (pH 8), which is similar to previous research.51 Fe(II) appeared to slightly enhance the adsorption of PHA. Previous work also reported that Fe(II) can increase oxyanion adsorption onto iron oxides, such as sulfate and phosphate, through forming ternary surface complexes.26 The adsorption of metal ion (Me)−ligand (L) complexes onto metal oxide surfaces (S) may proceed through three different mechanisms: (1) the adsorption of aqueous Me−L complexes by electrostatic forces or hydrogen bonding onto the metal oxide surface,71 (2) type A ternary complex formation where the metal ion bridges the metal oxide surface and the ligand (S−Me−L),31 and (3) type B ternary complex formation where the ligand is between the metal oxide surface and the metal ion (S−L−Me).72 For example, it has been suggested that the Cu(II)−glyphosate complex adsorbed on goethite was type A at high pH but type B at low pH.73 A number of ternary systems containing PHA, non-Fe(II) metal ions, and iron oxides have been studied.30,52,53,74,75 In most of these studies, type A species were believed to dominate, because steric constraints will inhibit the formation of type B species on iron oxide surfaces.53 Therefore, in our Fe(II)−PHA−goethite systems, we also believe that if there were ternary species formed, type A species were more likely than type B. Surface Complexation Modeling. SCM can offer additional information to clarify what species formed on the goethite surface when Fe(II) and PHA were present. In the binary system of Fe(II)/goethite, there were mainly two species (FeOFe+ and FeOFeOH) formed on the goethite surface (Figure S5), which is similar to previous studies.11,17 The Fe(II) monohydroxo surface complex (FeOFeOH) was believed to be the reactive species that reduced different contaminants due to its lower reduction potential.13 When PHA was present, there were three Fe(II) surface species formed according to the modeling data (Figure 4a). Besides the two binary species discussed above, there was one additional outer-sphere ternary surface species (( FeOFe+)2···L2−). The estimated amount of the ternary species increased gradually to maximum and then decreased as pH increased. In the binary system of PHA/goethite, the concentration of the outer-sphere binary species ((FeOH2+)2···L2−) gradually decreased as pH increased, whereas that of the inner-sphere species (Fe2L) stayed low (Figure S6), which is consistent

shown that aqueous Fe(II) was oxidized and underwent atom exchange upon adsorption onto Fe(III) oxide surfaces,68−70 aqueous Fe(II) was reported to behave macroscopically as an adsorbing divalent cation.26 As shown in Figure 3a, the

Figure 3. (a) Adsorption of 250 μM Fe(II) onto goethite (5 g/L) in the presence of three different concentrations of PHA (0, 50, and 200 μM) under different pHs; (b) adsorption of PHA (50 μM) onto goethite (5 g/L) in the absence and presence of 250 μM Fe(II) under different pHs. The adsorption percentage (P) was calculated based on P = (C0 − C)/C0 × 100%. Reaction conditions: 0.1 M NaCl, 20 mM MES buffer for pH 5.5−6.5, and 20 mM MOPs buffer for pH 7−8.

adsorption of Fe(II) onto goethite was pH-dependent and increased as pH increased. However, unlike the studies reporting that sulfate and phosphate slightly enhanced the adsorption of Fe(II) onto iron oxides,26 the adsorption of Fe(II) was not significantly affected by PHA. The adsorption of PHA to goethite as a function of pH is shown in Figure 3b.

Figure 4. Optimized TLM for Fe(II) and PHA adsorption in the ternary systems (250 μM Fe(II), 50 μM PHA, and 5 g/L goethite). Points are the experimental data and lines are the model fits. The solid lines are the overall model fits, and the dashed lines represent the fractional contributions of each species. 8224

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227

Article

Langmuir

Despite many years’ research, the detailed mechanism about how electrons transfer from Fe(II) to iron minerals and then to contaminants is still not clear. Based on our experimental and modeling data and Gorski and Scherer,77 it is possible that the adsorption of aqueous Fe(II) and the subsequent electron transfer between the adsorbed Fe(II) and the conduction band of goethite are two separate reactions, although related. This view may serve as a bridge between the classical view of adsorbed Fe(II) as a stable chemical reductant and the recent view that electrons transfer between the adsorbed Fe(II) and the contaminants through the conduction band of the Fe oxides.77 It is likely that once Fe(II) has been adsorbed, different adsorbed Fe(II) species form, and the rate of electron transfer from the adsorbed Fe(II) species to the contaminant depends on the nature of the formed species, for example, reduction potential and the reactivity of the site. In the presence of PHA, the Fe(II) adsorption and formation of adsorbed species may be affected, which then alters the reductive reactivity of the Fe(II)/goethite system. Future understanding of other ternary mixtures will be important to predict the reactivity of the mixtures and a number of related biogeochemical processes.

with previous studies reporting that inner-sphere species can only form at lower pH.49,76 In the ternary system of Fe(II)/ PHA/goethite (Figure 4b), part of (FeOH2+)2···L2− was converted to (FeOFe+)2···L2− at higher pH. The fraction of the ternary species increased to a maximum and then gradually decreased (Figure 4b). In our ternary systems, the adsorbed PHA might change the types of adsorbed Fe(II) species, which decreased the reactivity. When PHA was present, the distribution of surface Fe(II) species was different under different pH conditions. As shown in Figure 4a, FeOFe+ was the dominant surface complex under different pHs, which was similar to that in the binary system (Figure S5). However, at pH 6, the abundance of the species followed FeOFe+ > (FeOFe+)2···L2− >  FeOFeOH, whereas at pH 7, FeOFe+ > FeOFeOH > ( FeOFe+)2···L2−. The above results are mainly due to the difference in PHA adsorption under different pHs. The extent of PHA adsorption was much higher at pH 6 than at pH 7 (Figure 3b), which resulted in the conversion of more  FeOFe+ and FeOFeOH into (FeOFe+)2···L2− at pH 6 than at pH 7. Previous research observed that the amount of FeOFeOH in binary Fe(II)/Fe(III) oxides linearly correlated with the reactivity (k) and that the k value approximately doubled when the concentration of FeOFeOH doubled.11,17 As the results shown in Table S2 for the ternary mixture, the abundance of FeOFeOH and FeOFe+ decreased from 54 to 51.9% and from 6.01 to 5.71%, respectively, while that of (FeOFe+)2··· L2− increased to 8.87%. These results indicated that the conversion of the reactive binary species to the ternary species might have inhibited the electron transfer between the adsorbed Fe(II) and pCNB. It has been postulated that humic substances might have changed the type of reactive surface species formed, resulting in the decrease in the reductive reactivity.27 Now our results provided direct evidence to support this notion. However, based on the modeling data, the abundance of  FeOFeOH and FeOFe+ only decreased by 5 and 3.9%, respectively, at pH 6 when 50 μM PHA was added (Table S2), yet the k value decreased drastically by 56.4% (Figure 1). Therefore, it is likely that PHA had inhibited the reductive reactivity through two additional mechanisms: (1) Iron minerals are known to have different surface sites that have different reactivities.3 The adsorption of smaller amounts of PHA on goethite surfaces might have blocked the involvement of the more reactive sites in pCNB reduction, so there was a sharp decrease in k with increasing PHA concentration up to ∼50 μM at pH 6 (Figure 1). With large amounts of PHA (initial concentration > 50 μM) adsorbed on goethite surfaces, the decrease in k became much slower with increasing PHA concentration because the (re)generation of reactive surface sites might be rate-limiting.10,16 (2) Gorski and Scherer have demonstrated that the electrons donated from Fe(II) can go through the bulk conduction band of iron minerals and then reacted with different contaminants in the aqueous phase.77 Here, the PHA adsorbed onto goethite surfaces might have blocked this path. This is similar to previous research proposing that the adsorption of humic acid might interfere with electron transfer from adsorbed Fe(II) to the conduction band and then to organic contaminants.78 However, future research is needed to confirm these two mechanisms.

■

CONCLUSIONS This work demonstrated the effect of a LMW carboxylic acid on the reductive reactivity of Fe(II)/goethite, which offers mechanistic insights into the surface speciation and the corresponding impact on the reduction reactivity of Fe(II)/ goethite containing PHA. PHA significantly inhibited the reductive reactivity of Fe(II)/goethite by three mechanisms: (1) changing the adsorbed Fe(II) species from binary to ternary species, which decreased the abundance of the reactive Fe(II) species, (2) blocking the more reactive sites, and (3) inhibiting the electron transfer from Fe(II) to goethite and then to pCNB. The important role of surface Fe(II) speciation was demonstrated in ternary mixtures. Such information is crucial in the understanding and modeling of the fate of reducible contaminants in complex environments.

■

ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.langmuir.9b00538. Properties of goethite; distribution of Fe(II) species based on the modeling data; generic pCNB reduction kinetics; effect of PHA concentration on the k values at pH 7; the adsorption of Fe(II) and PHA; sedimentation kinetics; speciation of PHA at different pHs; and TLM for Fe(II) and PHA adsorption onto goethite (PDF)

■

AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. ORCID

Zimeng Wang: 0000-0002-4572-629X Huichun Zhang: 0000-0002-5683-5117 Notes

The authors declare no competing financial interest. 8225

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227

Article

Langmuir

■

(18) Amirbahman, A.; Kent, D. B.; Curtis, G. P.; Marvin-DiPasquale, M. C. Kinetics of Homogeneous and Surface-Catalyzed Mercury(II) Reduction by Iron(II). Environ. Sci. Technol. 2013, 47, 7204−7213. (19) Huang, J.; Dai, Y.; Liu, C.-C.; Zhang, H. Effects of Second Metal Oxides on Surface-Mediated Reduction of Contaminants by Fe(II) with Iron Oxide. ACS Earth Space Chem. 2019, 3, 680−687. (20) Strathmann, T. J.; Stone, A. T. Mineral surface catalysis of reactions between FeII and oxime carbamate pesticides. Geochim. Cosmochim. Acta 2003, 67, 2775−2791. (21) Elsner, M.; Schwarzenbach, R. P.; Haderlein, S. B. Reactivity of Fe(II)-bearing minerals toward reductive transformation of organic contaminants. Environ. Sci. Technol. 2004, 38, 799−807. (22) Vikesland, P. J.; Valentine, R. L. Iron oxide surface-catalyzed oxidation of ferrous iron by monochloramine: Implications of oxide type and carbonate on reactivity. Environ. Sci. Technol. 2002, 36, 512− 519. (23) Chun, C. L.; Penn, R. L.; Arnold, W. A. Kinetic and microscopic studies of reductive transformations of organic contaminants on goethite. Environ. Sci. Technol. 2006, 40, 3299−3304. (24) Latta, D. E.; Bachman, J. E.; Scherer, M. M. Fe electron transfer and atom exchange in goethite: influence of Al-substitution and anion sorption. Environ. Sci. Technol. 2012, 46, 10614−10623. (25) Jones, A. M.; Collins, R. N.; Rose, J.; Waite, T. D. The effect of silica and natural organic matter on the Fe(II)-catalysed transformation and reactivity of Fe(III) minerals. Geochim. Cosmochim. Acta 2009, 73, 4409−4422. (26) Hinkle, M. A. G.; Wang, Z.; Giammar, D. E.; Catalano, J. G. Interaction of Fe(II) with phosphate and sulfate on iron oxide surfaces. Geochim. Cosmochim. Acta 2015, 158, 130−146. (27) Vindedahl, A. M.; Stemig, M. S.; Arnold, W. A.; Penn, R. L. Character of humic substances as a predictor for goethite nanoparticle reactivity and aggregation. Environ. Sci. Technol. 2016, 50, 1200− 1208. (28) Colón, D.; Weber, E. J.; Anderson, J. L. Effect of natural organic matter on the reduction of nitroaromatics by Fe(II) species. Environ. Sci. Technol. 2008, 42, 6538−6543. (29) Benjamin, M. M.; Leckie, J. O. Effects of complexation by chloride, sulfate, and thiosulfate on adsorption behavior of cadmium on oxide surfaces. Environ. Sci. Technol. 1982, 16, 162−170. (30) Ali, M. A.; Dzombak, D. A. Effects of simple organic acids on sorption of Cu2+ and Ca2+on goethite. Geochim. Cosmochim. Acta 1996, 60, 291−304. (31) Ostergren, J. D.; Brown, G. E., Jr; Parks, G. A.; Persson, P. Inorganic Ligand Effects on Pb(II) Sorption to Goethite (α-FeOOH). J. Colloid Interface Sci. 2000, 225, 483−493. (32) Swedlund, P. J.; Webster, J. G.; Miskelly, G. M. The effect of SO4 on the ferrihydrite adsorption of Co, Pb and Cd: ternary complexes and site heterogeneity. Appl. Geochem. 2003, 18, 1671− 1689. (33) Swedlund, P. J.; Webster, J. G.; Miskelly, G. M. Goethite adsorption of Cu(II), Pb(II), Cd(II), and Zn(II) in the presence of sulfate: Properties of the ternary complex. Geochim. Cosmochim. Acta 2009, 73, 1548−1562. (34) Zaman, M. I.; Mustafa, S.; Khan, S.; Xing, B. Effect of phosphate complexation on Cd2+ sorption by manganese dioxide (βMnO2). J. Colloid Interface Sci. 2009, 330, 9−19. (35) Ali, M. A.; Dzombak, D. A. Interactions of copper, organic acids, and sulfate in goethite suspensions. Geochim. Cosmochim. Acta 1996, 60, 5045−5053. (36) Juang, R.-S.; Wu, W.-L. Adsorption of sulfate and copper(II) on goethite in relation to the changes of zeta potentials. J. Colloid Interface Sci. 2002, 249, 22−29. (37) Jeon, B.-H.; Dempsey, B. A.; Burgos, W. D.; Royer, R. A. Reactions of ferrous iron with hematite. Colloids Surf. A 2001, 191, 41−55. (38) Swedlund, P. J.; Webster, J. G. Cu and Zn ternary surface complex formation with SO4 on ferrihydrite and schwertmannite. Appl. Geochem. 2001, 16, 503−511.

ACKNOWLEDGMENTS This material is based upon work supported by the National Science Foundation under Grants CBET-1762691 and CHE1762686 and National Natural Science Foundation of China (21806021). The authors are thankful to Dr. Andro-Marc Pierre-Louis at Temple University for the assistance in using ATR−FTIR.

■

REFERENCES

(1) Orsetti, S.; Laskov, C.; Haderlein, S. B. Electron transfer between iron minerals and quinones: estimating the reduction potential of the Fe(II)-goethite surface from AQDS speciation. Environ. Sci. Technol. 2013, 47, 14161−14168. (2) Wang, Z.; Schenkeveld, W. D. C.; Kraemer, S. M.; Giammar, D. E. Synergistic Effect of Reductive and Ligand-Promoted Dissolution of Goethite. Environ. Sci. Technol. 2015, 49, 7236−7244. (3) Chen, Y.; Dong, H.; Zhang, H. Experimental and Computational Evidence for the Reduction Mechanisms of Aromatic N-oxides by Aqueous FeII-Tiron Complex. Environ. Sci. Technol. 2016, 50, 249− 258. (4) Chen, Y.; Zhang, H. Complexation Facilitated Reduction of Aromatic N-Oxides by Aqueous FeII-Tiron Complex: Reaction Kinetics and Mechanisms. Environ. Sci. Technol. 2013, 47, 11023− 11031. (5) Zhang, H.; Weber, E. J. Identifying indicators of reactivity for chemical reductants in sediments. Environ. Sci. Technol. 2012, 47, 6959−6968. (6) Zhang, H.; Weber, E. J. Elucidating the role of electron shuttles in reductive transformations in anaerobic sediments. Environ. Sci. Technol. 2009, 43, 1042−1048. (7) Chen, Y.; Dong, H.; Zhang, H. Reduction of isoxazoles including sulfamethoxazole by aqueous FeII-tiron complex: Impact of structures. Chem. Eng. J. 2018, 352, 501−509. (8) Li, X.; Chen, Y.; Zhang, H. Reduction of nitrogen-oxygen containing compounds (NOCs) by surface-associated Fe(II) and comparison with soluble Fe(II) complexes. Chem. Eng. J. 2019, 370, 782−791. (9) Pecher, K.; Haderlein, S. B.; Schwarzenbach, R. P. Reduction of Polyhalogenated Methanes by Surface-Bound Fe(II) in Aqueous Suspensions of Iron Oxides. Environ. Sci. Technol. 2002, 36, 1734− 1741. (10) Klausen, J.; Troeber, S. P.; Haderlein, S. B.; Schwarzenbach, R. P. Reduction of Substituted Nitrobenzenes by Fe(II) in Aqueous Mineral Suspensions. Environ. Sci. Technol. 1995, 29, 2396−2404. (11) Charlet, L.; Silvester, E.; Liger, E. N-compound reduction and actinide immobilisation in surficial fluids by Fe(II): the surface FeIIIOFeIIOH° species as major reductant. Chem. Geol. 1998, 151, 85−93. (12) Klupinski, T. P.; Chin, Y.-P.; Traina, S. J. Abiotic degradation of pentachloronitrobenzene by Fe(II): reactions on goethite and iron oxide nanoparticles. Environ. Sci. Technol. 2004, 38, 4353−4360. (13) Silvester, E.; Charlet, L.; Tournassat, C.; Géhin, A.; Grenèche, J.-M.; Liger, E. Redox potential measurements and Mössbauer spectrometry of FeII adsorbed onto FeIII (oxyhydr)oxides. Geochim. Cosmochim. Acta 2005, 69, 4801−4815. (14) Wang, S.; Arnold, W. A. Abiotic reduction of dinitroaniline herbicides. Water Res. 2003, 37, 4191−4201. (15) Klupinski, T. P.; Chin, Y.-P. Abiotic degradation of trifluralin by Fe(II): kinetics and transformation pathways. Environ. Sci. Technol. 2003, 37, 1311−1318. (16) Amonette, J. E.; Workman, D. J.; Kennedy, D. W.; Fruchter, J. S.; Gorby, Y. A. Dechlorination of carbon tetrachloride by Fe(II) associated with goethite. Environ. Sci. Technol. 2000, 34, 4606−4613. (17) Liger, E.; Charlet, L.; Van Cappellen, P. Surface catalysis of uranium(VI) reduction by iron(II). Geochim. Cosmochim. Acta 1999, 63, 2939−2955. 8226

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227

Article

Langmuir (39) Girvin, D. C.; Gassman, P.; Bolton, H. Adsorption of nitrilotriacetate (NTA), Co and CoNTA by gibbsite. Clays Clay Miner. 1996, 44, 757−768. (40) Buerge-Weirich, D.; Hari, R.; Xue, H.; Behra, P.; Sigg, L. Adsorption of Cu, Cd, and Ni on goethite in the presence of natural groundwater ligands. Environ. Sci. Technol. 2002, 36, 328−336. (41) Elzinga, E. J.; Kretzschmar, R. In situ ATR-FTIR spectroscopic analysis of the co-adsorption of orthophosphate and Cd(II) onto hematite. Geochim. Cosmochim. Acta 2013, 117, 53−64. (42) Strobel, B. W. Influence of vegetation on low-molecular-weight carboxylic acids in soil solution-a review. Geoderma 2001, 99, 169− 198. (43) Axe, K.; Persson, P. Time-dependent surface speciation of oxalate at the water-boehmite (γ-AlOOH) interface: implications for dissolution. Geochim. Cosmochim. Acta 2001, 65, 4481−4492. (44) Jonsson, S.; Ejlertsson, J.; Ledin, A.; Mersiowsky, I.; Svensson, B. H. Mono- and diesters from o-phthalic acid in leachates from different European landfills. Water Res. 2003, 37, 609−617. (45) Banwart, S.; Davies, S.; Stumm, W. The role of oxalate in accelerating the reductive dissolution of hematite (α-Fe2O3) by ascorbate. Colloids Surf. 1989, 39, 303−309. (46) Stumm, W. Reactivity at the mineral-water interface: dissolution and inhibition. Colloids Surf. A 1997, 120, 143−166. (47) Biber, M. V.; dos Santos Afonso, M.; Stumm, W. The coordination chemistry of weathering: IV. Inhibition of the dissolution of oxide minerals. Geochim. Cosmochim. Acta 1994, 58, 1999−2010. (48) Taujale, S.; Baratta, L. R.; Huang, J.; Zhang, H. Interactions in ternary mixtures of MnO2, Al2O3, and natural organic matter (NOM) and the impact on MnO2 oxidative reactivity. Environ. Sci. Technol. 2016, 50, 2345−2353. (49) Zhang, H.; Taujale, S.; Huang, J.; Lee, G.-J. Effects of NOM on oxidative reactivity of manganese dioxide in binary oxide mixtures with goethite or hematite. Langmuir 2015, 31, 2790−2799. (50) Gu, B.; Schmitt, J.; Chen, Z.; Liang, L.; McCarthy, J. F. Adsorption and desorption of different organic matter fractions on iron oxide. Geochim. Cosmochim. Acta 1995, 59, 219−229. (51) Hwang, Y. S.; Liu, J.; Lenhart, J. J.; Hadad, C. M. Surface complexes of phthalic acid at the hematite/water interface. J. Colloid Interface Sci. 2007, 307, 124−134. (52) Song, Y.; Swedlund, P. J.; Singhal, N. Copper(II) and cadmium(II) sorption onto ferrihydrite in the presence of phthalic acid: some properties of the ternary complex. Environ. Sci. Technol. 2008, 42, 4008−4013. (53) Boily, J.-F.; Sjöberg, S.; Persson, P. Structures and stabilities of Cd(II) and Cd(II)-phthalate complexes at the goethite/water interface. Geochim. Cosmochim. Acta 2005, 69, 3219−3235. (54) Boily, J.-F.; Persson, P.; Sjöberg, S. Benzenecarboxylate surface complexation at the goethite (α-FeOOH)/water interface: II. Linking IR spectroscopic observations to mechanistic surface complexation models for phthalate, trimellitate, and pyromellitate. Geochim. Cosmochim. Acta 2000, 64, 3453−3470. (55) Colón, D.; Weber, E. J.; Baughman, G. L. Sediment-associated reactions of aromatic amines. 2. QSAR development. Environ. Sci. Technol. 2002, 36, 2443−2450. (56) Colón, D.; Weber, E. J.; Anderson, J. L.; Winget, P.; Suárez, L. A. Reduction of Nitrosobenzenes andN-Hydroxylanilines by Fe(II) Species: Elucidation of the Reaction Mechanism. Environ. Sci. Technol. 2006, 40, 4449−4454. (57) Cwiertny, D. M.; Handler, R. M.; Schaefer, M. V.; Grassian, V. H.; Scherer, M. M. Interpreting nanoscale size-effects in aggregated Fe-oxide suspensions: Reaction of Fe(II) with Goethite. Geochim. Cosmochim. Acta 2008, 72, 1365−1380. (58) Viollier, E.; Inglett, P. W.; Hunter, K.; Roychoudhury, A. N.; Van Cappellen, P. The ferrozine method revisited: Fe(II)/Fe(III) determination in natural waters. Applied Geochem 2000, 15, 785−790. (59) Yang, Y.; Duan, J.; Jing, C. Molecular-scale study of salicylate adsorption and competition with catechol at goethite/aqueous solution interface. J. Phys. Chem. C 2013, 117, 10597−10606.

(60) Xie, X.; Giammar, D. E.; Wang, Z. MINFIT: A SpreadsheetBased Tool for Parameter Estimation in an Equilibrium Speciation Software Program. Environ. Sci. Technol. 2016, 50, 11112−11120. (61) Wang, Z.; Giammar, D. E. Mass Action Expressions for Bidentate Adsorption in Surface Complexation Modeling: Theory and Practice. Environ. Sci. Technol. 2013, 47, 3982−3996. (62) Park, C. M. Mercury speciation during thermal remediation and in post-treatment environments. Doctoral Dissertation, University of Texas at Austin, 2011. (63) Hwang, Y. S.; Lenhart, J. J. Surface complexation modeling of dual-mode adsorption of organic acids: Phthalic acid adsorption onto hematite. J. Colloid Interface Sci. 2009, 336, 200−207. (64) Nano, G. V.; Strathmann, T. J. Application of surface complexation modeling to the reactivity of iron(II) with nitroaromatic and oxime carbamate contaminants in aqueous TiO2 suspensions. J. Colloid Interface Sci. 2008, 321, 350−359. (65) Bonnissel-Gissinger, P.; Alnot, M.; Lickes, J.-P.; Ehrhardt, J.-J.; Behra, P. Modeling the Adsorption of Mercury(II) on (Hydr)oxides II: α-FeOOH (Goethite) and Amorphous Silica. J. Colloid Interface Sci. 1999, 215, 313−322. (66) Vikesland, P. J.; Heathcock, A. M.; Rebodos, R. L.; Makus, K. E. Particle size and aggregation effects on magnetite reactivity toward carbon tetrachloride. Environ. Sci. Technol. 2007, 41, 5277−5283. (67) Yang, Y.; Du, J.; Jing, C. Dynamic adsorption process of phthalate at goethite/aqueous interface: An ATR-FTIR study. Colloids Surf. A 2014, 441, 504−509. (68) Williams, A. G. B.; Scherer, M. M. Spectroscopic Evidence for Fe(II)−Fe(III) Electron Transfer at the Iron Oxide−Water Interface. Environ. Sci. Technol. 2004, 38, 4782−4790. (69) Handler, R. M.; Beard, B. L.; Johnson, C. M.; Scherer, M. M. Atom Exchange between Aqueous Fe(II) and Goethite: An Fe Isotope Tracer Study. Environ. Sci. Technol. 2009, 43, 1102−1107. (70) Crosby, H. A.; Johnson, C. M.; Roden, E. E.; Beard, B. L. Coupled Fe(II)−Fe(III) Electron and Atom Exchange as a Mechanism for Fe Isotope Fractionation during Dissimilatory Iron Oxide Reduction. Environ. Sci. Technol. 2005, 39, 6698−6704. (71) Bargar, J. R.; Persson, P.; Brown, G. E., Jr Outer-sphere adsorption of Pb(II)EDTA on goethite. Geochim. Cosmochim. Acta 1999, 63, 2957−2969. (72) Wang, Y.-J.; Zhou, D.-M.; Sun, R.-J.; Jia, D.-A.; Zhu, H.-W.; Wang, S.-Q. Zinc adsorption on goethite as affected by glyphosate. J. Hazard. Mater. 2008, 151, 179−184. (73) Sheals, J.; Granström, M.; Sjöberg, S.; Persson, P. Coadsorption of Cu(II) and glyphosate at the water-goethite (α-FeOOH) interface: molecular structures from FTIR and EXAFS measurements. J. Colloid Interface Sci. 2003, 262, 38−47. (74) Lövgren, L. Complexation reactions of phthalic acid and aluminium (III) with the surface of goethite. Geochim. Cosmochim. Acta 1991, 55, 3639−3645. (75) Angove, M. J.; Wells, J. D.; Johnson, B. B. Adsorption of cadmium(II) onto goethite and kaolinite in the presence of benzene carboxylic acids. Colloids Surf. A 1999, 146, 243−251. (76) Boily, J.-F.; Nilsson, N.; Persson, P.; Sjöberg, S. Benzenecarboxylate Surface Complexation at the Goethite (α-FeOOH)/ Water Interface: I. A Mechanistic Description of Pyromellitate Surface Complexes from the Combined Evidence of Infrared Spectroscopy, Potentiometry, Adsorption Data, and Surface Complexation Modeling. Langmuir 2000, 16, 5719−5729. (77) Gorski, C.; Scherer, M., Fe2+ sorption at the Fe oxide-water interface: A revised conceptual framework. Aquatic Redox Chemistry. ACS Symposium Series, 2011; Vol. 1071, pp 314−343. (78) Luan, F.; Xie, L.; Li, J.; Zhou, Q. Abiotic reduction of nitroaromatic compounds by Fe(II) associated with iron oxides and humic acid. Chemosphere 2013, 91, 1035−1041.

8227

DOI: 10.1021/acs.langmuir.9b00538 Langmuir 2019, 35, 8220−8227