Water, Air, & Soil Pollution

, Volume 216, Issue 1, pp 473–483

Influences of Humic Acid on Cr(VI) Removal by Zero-Valent Iron From Groundwater with Various Constituents: Implication for Long-Term PRB Performance

Authors

  • Tongzhou Liu
    • Department of Civil and Environmental EngineeringThe Hong Kong University of Science and Technology
    • AECOM (Hong Kong) Ltd
    • Department of Civil and Environmental EngineeringThe Hong Kong University of Science and Technology
Article

DOI: 10.1007/s11270-010-0546-2

Cite this article as:
Liu, T. & Lo, I.M.C. Water Air Soil Pollut (2011) 216: 473. doi:10.1007/s11270-010-0546-2

Abstract

A 9-month-long continuous flow column study was carried out to investigate Cr(VI) removal by Fe0 with the presence of humic acid. The study focused on the influences of humic acid promoted dissolved iron release and humic acid aggregation in Fe0 columns receiving synthetic Cr(VI) contaminated groundwater containing various components such as bicarbonate and Ca. The effects of humic acid varied significantly depending on the presence of Ca. In Ca-free columns, the presence of humic acid promoted the release of dissolved iron in the forms of soluble Fe-humic acid complexes and stabilized fine Fe (hydr)oxide colloids. As a result, the precipitation of iron corrosion products was suppressed and the accumulation of secondary minerals on Fe0 surfaces was diminished, and a slight increase in Cr(VI) removal capacity by 18% was record compared with that of humic acid-free column. In contrast, in the presence of Ca, as evidenced by the SEM and FTIR results, humic acid greatly co-aggregated with Fe (hydr)oxides and deposited on Fe0 surfaces. This largely inhibited electron transfer from Fe0 surfaces to Cr(VI) and reduced the drainable porosity of the Fe0 matrix, resulting in a significant decrease in Cr(VI) removal capacity of Fe0. The Cr(VI) removal capacity was decreased by 24.4% and 42.7% in humic acid and Ca receiving columns, with and without bicarbonate respectively, compared with that of Ca and humic acid-free column. This study yields new considerations for the performance prediction and design of Fe0 PRBs in the environments rich in natural organic matter (NOM).

Keywords

Column experimentsHumic acidPermeable reactive barrierZero-valent iron

1 Introduction

Zero-valent iron (Fe0) based permeable reactive barriers (PRBs) have exhibited great potential for treating a number of inorganic and organic contaminants, such as Cr(VI), U(VI), chlorinated hydrocarbons, and nitroaromatic compounds, etc. (Gillham and O’Hannesin 1994; Cantrell et al. 1995; Powell et al. 1995; Klausen et al. 2003). However, uncertainties remain concerning the long-term effectiveness of Fe0 PRBs in complex subsurface environments. The continuous buildup of secondary minerals on Fe0 surfaces, including amorphous and crystalline iron (hydr)oxides, carbonate minerals, as well as iron sulfides (Gu et al. 1999; Roh et al. 2000; Lo et al. 2006) is believed to negatively affect the long-term effectiveness of Fe0 PRBs by inhibiting the access of contaminants to the Fe0 surfaces (Roh et al. 2000; Klausen et al. 2003; Lo et al. 2006) and deteriorating the hydraulic performance of the Fe0 matrix over time (Kamolpornwijit et al. 2003; Wilkin et al. 2005).

Besides the secondary minerals, natural organic matter (NOM) may also influence the performance of Fe0 PRBs. In spite of the reducing capacity associated with some minor electron shuttling functional groups, such as quinones (Tratnyek et al. 2001), NOM was not observed to play an important role in reducing contaminants in engineered systems with Fe0 as a predominant reductant (Tratnyek et al. 2001; Xie and Shang 2005). On the other hand, more attention was focused on the complexation of NOM with metals, as well as the competition for surface sorption sites between NOM and contaminants. Dries et al. (2005) indicated that the formation of metal-humic acid complexes in solution was responsible for the delayed removal of Zn2+ and Ni2+ by Fe0. Humic and fulvic acids were also reported to have inhibitory effects on the reductive degradation of chlorinated and nitroaromatic organics by Fe0, which was ascribed to the competition for reactive sites between adsorbed humic acid and organic contaminants (Tratnyek et al. 2001; Klausen et al. 2003).

In addition to the sorption of NOM onto Fe0 surfaces and the complexation with heavy metals, recent studies also suggested occurrences of the NOM promoted release of “dissolved” iron and the NOM aggregation in Fe0 systems, depending on solution conditions (Dries et al. 2005; Liu et al. 2008). The “dissolved” iron (as operationally defined by filtering through 0.45-μm membrane), probably comprising soluble Fe-NOM complexes and stable fine Fe (hydr)oxide colloids (Tipping 2002), might suppress the accumulation of iron corrosion products on Fe0 surfaces and then favor the removal of contaminants by Fe0. Nevertheless, as observed in batch experiments, when hardness (Ca or Mg) existed, large aggregates comprising Fe (hydr)oxides and humic acid were formed (Liu et al. 2008), and were suspected of progressively depositing on Fe0 surfaces and thus adversely affect the long-term performance of Fe0 PRBs. The effects of these two phenomena, however, remain speculative. In order to clarify this speculation and further assess the effects introduced by NOM, and also to provide engineers with additional information for designing Fe0 PRBs in NOM-rich environments, studies are worth carrying out in conditions that can mimic the flow patterns and solid to liquid ratio in a real Fe0 PRB. This study, therefore, investigated the influences of humic acid on Cr(VI) removal by Fe0 in a series of continuous flow columns. The columns were packed with Fe0 and fed with synthetic Cr(VI) contaminated groundwater containing various components, including humic acid, Ca, and bicarbonate. Focus was directed on the occurrences of humic acid promoted dissolved iron release and humic acid aggregation, under different geochemical conditions, and their effects on Cr(VI) removal by Fe0.

2 Experimental Section

2.1 Materials

Sieved fractions (0.5–1 mm) of granular zero-valent iron (Fe0; ETI-CC-1004, Connelly-GPM Inc., U. S.) were used in this study. Chemical stock solutions were prepared by dissolving reagent-grade chemicals, K2Cr2O7, NaCl, CaCl2·2H2O, NaHCO3, NaOH, and HCl (Riedel-de Haën), into ultrapure water (Barnstead D11911). Humic acid stock solution was prepared by dissolving a certain amount of humic acid sodium salt (Aldrich) into ultrapure water followed by filtering through 0.45-μm acetate cellulose membranes (ADVANTEC). The humic acid concentration was measured using a combustion-infrared TOC analyzer (Shimadzu TOC 5000A) and expressed as dissolved organic carbon (DOC). Synthetic groundwater with Cr(VI) and various components, such as humic acid, bicarbonate and Ca, was prepared by diluting the chemical stock solutions to target concentrations using ultrapure water. The initial pH of the prepared synthetic groundwater was adjusted to 7.0 ± 0.1 by adding 0.01 N NaOH and 0.01 N HCl for batch experiments and 0.1 N NaOH and 0.1 N HCl for column tests. The concentrations of the components and the initial pH in solutions are within their typical ranges in groundwater (Snoeyink and Jenkins 1980; Thurman 1985).

2.2 Column and Batch Experiments

Columns used in this study were made of rigid PVC pipe, measuring 30 cm in length by 3.6 cm in internal diameter. Sampling ports were arranged along the column every 5 cm from the influent end (see Supporting Information Fig. S1). Granular Fe0 was carefully packed in the columns to achieve a homogeneous distribution with an approximate bulk density of 3.08 g cm−3 and a calculated porosity of 0.52. Eight columns were set up to investigate the effects of humic acid, Ca and bicarbonate on Cr(VI) removal by Fe0, individually and collectively (see Table 1). Synthetic groundwater was fed in an up-flow manner at a groundwater velocity of 400 m year−1 (1.29 × 10−3 cm s−1) using a multi-channel precision peristaltic pump (MasterFlex). This value is within the range of groundwater velocity (21.9–803 m year−1/6.94 × 10−5–2.55 × 10−3 cm s−1) reported for PRBs in the field (Klausen et al. 2003) and also applicable for PRBs with a funnel-and-gate configuration, where groundwater velocities are greater than those resulting from the natural gradient (USEPA 1998). The required time for every pore volume (PV) was about 6.6–6.7 h. Water samples (around 5 ml) were collected regularly (every 20–30 PV) from the sampling ports as well as both the influent and effluent ends of the columns for Cr(VI) concentration measurement. For some parameters of less frequent measurements such as pH, concentrations of humic acid, total dissolved iron, Ca, and total alkalinity, 40 ml of water samples were collected from the influent and effluent ends every 40–50 PV. Water samples were filtered through 0.45-μm membranes at the time of collection and analyzed within 24 h after collection. At the end of the column experiments, an estimation of the amount of precipitates in different columns was conducted based on the chemical monitoring data (Supporting Information Table S1).
Table 1

The composition of the synthetic groundwater, Cr(VI) front migration rate, Cr(VI) removal capacity of Fe0, and effective porosity in different columns

Columna

Cr(VI) (mg l−1)

NaCl (mM)

Humic acid (mg l−1 DOC)

CaCl2 (mM)

NaHCO3 (mM)

Cr(VI) front migration rate (cm PV−1)b

Cr(VI) removal capacity of Fe0 (mg Cr(VI)/g Fe0)c

Drainable porosity at the end of the 9-month column experiment (%)d

0e

0

0

0

0

0

39.3

1

25

5

0

0

0

0.0306 ± 0.0012

4.15 ± 0.16

31.0

2

25

5

20

0

0

0.0259 ± 0.0006

4.89 ± 0.11

34.9

3

25

5

0

0.8

0

0.0354 ± 0.0012

3.58 ± 0.13

29.4

4

25

5

20

0.8

0

0.0396 ± 0.0017

3.20 ± 0.14

21.2

5

25

5

0

0

6

0.0267 ± 0.0009

4.76 ± 0.15

31.5

6

25

5

20

0

6

0.0332 ± 0.0007

3.81 ± 0.08

34.8

7

25

5

0

0.8

6

0.0441 ± 0.0018

2.88 ± 0.12

28.5

8

25

5

20

0.8

6

0.0531 ± 0.0040

2.40 ± 0.18

22.9

aThe columns were 30 cm in length. Synthetic groundwater was fed in an up-flow manner at a groundwater velocity of 400 m year−1 (1.29 × 10−3 cm s−1).

bThe errors indicate standard deviation of Cr(VI) front migration rates at relative Cr(VI) concentrations of 0.3, 0.5, and 0.7.

cThe errors indicate calculated standard deviation of Cr(VI) removal capacity based on Cr(VI) front migration rates at relative Cr(VI) concentrations of 0.3, 0.5, and 0.7.

dTo avoid the distortion of the columns when refilling them, the drainable porosity was measured only once at the end of the column experiment.

eColumn 0 was fed with ultrapure water for 1 week to obtain the drainable porosity without the introduction of groundwater solutes.

A batch experiment was carried out to evaluate the direct Cr(VI) reduction by humic acid. The experiment was conducted in three 250-ml plastic bottles containing 200 ml of solution with 2 mg l−1 of Cr(VI) and 20 mg l−1 DOC of humic acid. The initial solution pH in these bottles was adjusted from 5.2 to 10.3, respectively, by adding 0.01 N NaOH and 0.01 N HCl. Bottles were shaken end-over-end at 26 rpm at room temperature (23 ± 1°C) in dark. Solutions were sampled at regular time intervals up to 63 d, filtered through 0.45-μm membranes, and followed by immediate measurement of pH and the concentrations of total Cr and Cr(VI).

2.3 Analytical Methods

The Cr(VI) concentration was measured by the 1,5-diphenylcarbazide colorimetric method using a UV/visible spectrometer (Ultrospec 4300 Pro) at a wavelength of 540 nm. The pH was measured using a pH meter (Orion Model 420A) with a gel filled combination electrode. Concentrations of humic acid and dissolved metals (Cr, Fe, and Ca) were determined using a TOC analyzer (Shimadzu TOC 5000A), and an atomic absorption spectrometer (Varian 220FS), respectively. Total alkalinity was determined by titration using 0.01 N HCl as a titrant and mixed methyl red/bromocresol green solution as an indicator.

At the end of the column experiment, the drainable porosity of a column was calculated by draining the column, measuring the volume of fluid collected, and dividing this volume by the total volume within the column (VanGulck and Rowe 2004). It was applied to provide a useful indication of the level of clogging occurring in the columns that received different synthetic contaminated groundwater. Column 0 was set as a control column that was exposed to ultrapure water for 1 week. Its drainable porosity was regarded as the baseline of the porosity in columns and used to assess the level of clogging which occurred in columns that received synthetic contaminated groundwater. To avoid the distortion of column when refilling it, the drainable porosity in a column was measured only once. The column was then dismantled and the reacted Fe0 (about 10 g) within it was collected immediately. The collected reacted Fe0 was freeze-dried and then stored in sealed plastic bags purged with nitrogen gas. The surface as well as the cross-section of the reacted Fe0 were investigated using scanning electron microscopy coupled with energy dispersive X-ray (SEM-EDX, JEOL 6390). Cross-section samples were obtained by embedding the granular Fe0 in resins and subsequently grinding and polishing. X-ray diffraction analysis (XRD, Philips PW1825) was applied to identify the secondary minerals formed on the reacted Fe0 surfaces. In addition, the reacted Fe0 was ground by hand with a pestle and mortar to remove the surface coatings. The coatings were then collected and subject to Fourier transform infrared spectrometry (FTIR, PerkinElmer Spectrum BX) analysis.

3 Results and Discussion

3.1 Cr(VI) Removal

The removal mechanisms of Cr(VI) by Fe0 have been widely studied and believed to involve instantaneous adsorption of Cr(VI) on Fe0 surface where electron transfer takes place and Cr(VI) is reduced to Cr3+ with oxidation of Fe0 to Fe3+, and subsequently, Cr3+ precipitates as Cr3+ hydroxides and/or mixed Fe3+/Cr3+ (oxy)hydroxides (Eary and Rai 1988; Powell et al. 1995; Pratt et al. 1997; Blowes et al. 1997). The reactivity of Fe0 is highly dependent on the aqueous chemistry of groundwater. Column 1 was therefore set up as the reference column that received the Cr(VI) solution with 5 mM NaCl as the background electrolyte. Figure 1a shows the Cr(VI) breakthrough curves in column 1. The propagation of the Cr(VI) migration front was almost constant throughout the experiments (Fig. 1b). The same phenomenon was also observed in the remaining columns but with different migration rates, indicating stable Cr(VI) reaction and transportation process in different sections of the column. The slopes of the fitting lines in Fig. 1b are the Cr(VI) front migration rates (with the unit of cm PV−1) at relative Cr(VI) concentrations of 0.3, 0.5, and 0.7, respectively. The Cr(VI) removal capacity of Fe0 [mg Cr(VI)/g Fe0] can be estimated using the following equation:
$$ {\hbox{Removal capacity [mg Cr}}\left( {\hbox{VI}} \right)/{\hbox{g}}\,{\hbox{Fe] = }}\frac{{\left[ {{\text{Cr(VI)}}} \right]}}{{{\hbox{M}} \times {\hbox{A}} \times {\rho_{\rm{b}}} \times 1000}} $$
(1)
where [Cr(VI)] (mg l−1) is the initial concentration of Cr(VI), A (cm2) is the cross-sectional area of the column, ρb (g cm−3) is the bulk density of the Fe0 packed in the column, and M (cm cm−3) is the normalized migration rate of the Cr(VI) front at a relative Cr(VI) concentration of 0.3 or 0.5 or 0.7, which was calculated by dividing the migration rate (cm PV−1) by the pore volume of the column (cm3 PV−1). The average value and standard deviation of the Cr(VI) front migration rate and the Cr(VI) removal capacity at relative Cr(VI) concentrations of 0.3, 0.5, and 0.7, as well as the drainable porosities in different columns, are shown in Table 1.
https://static-content.springer.com/image/art%3A10.1007%2Fs11270-010-0546-2/MediaObjects/11270_2010_546_Fig1_HTML.gif
Fig. 1

a Cr(VI) breakthrough curves in column 1; b migration distance of the Cr(VI) front at C/C0 = 0.3, 0.5, or 0.7 in column 1, the slopes of the lines indicate the Cr(VI) front migration rates along the column. Column 1 received the Cr(VI) solution, with NaCl as the background electrolyte. The initial concentrations of Cr(VI) and NaCl were 25 mg l−1 and 5 mM, respectively

At the end of the column experiment, the Cr(VI) removal capacity in column 1 was 4.18 mg Cr(VI)/g Fe0 and the drainable porosity was determined to be 31.0% which was 8.3% lower than that of column 0 which was exposed to ultrapure water for 1 week. SEM images of the surface and the cross-section of the reacted Fe0 collected in column 1 revealed the presence of secondary minerals in euhedral tabular and irregular strip structures (Fig. 2a and b), indicating the formation of crystal Fe (hydr)oxides, such as Fe3O4, Fe(OH)3, and α, γ-FeOOH/α-Fe2O3 (Lai and Lo 2008). The accumulation rate of these Fe(hydr)oxides was estimated to be 7.53 mg/PV [as Fe(OH)3] throughout the column (Supporting Information Table S1). The buildup of Fe (hydr)oxides was believed to occlude the surface reactive sites and prevent electron transfer from the Fe0 surfaces to Cr(VI) (Roh et al. 2000), as well as to diminish the porosity of the Fe0 matrix over time (Kamolpornwijit et al. 2003; Wilkin et al. 2005). Both the passivation of Fe0 reactivity and the reduction of Fe0 matrix porosity would result in the depletion of the Cr(VI) removal capacity and finally lead to early Cr(VI) breakthrough.
https://static-content.springer.com/image/art%3A10.1007%2Fs11270-010-0546-2/MediaObjects/11270_2010_546_Fig2_HTML.gif
Fig. 2

SEM images of a and b the surface and the cross-section of the reacted Fe0 collected in column 1; c and d the surface and the cross-section of the reacted Fe0 collected in column 2; e and f the surface and the cross-section of the reacted Fe0 collected in column 4. The tables immediately below the images indicate the atomic composition of the solids in the encircled area by the EDX analysis

3.2 Effects of Ca and Bicarbonate

In column 3, which was exposed to the Cr(VI) contaminated synthetic groundwater with 0.8 mM of CaCl2, the Cr(VI) removal capacity was 3.51 mg Cr(VI)/g Fe0, and decreased by 16.1% compared with that of column 1 (Table 1). The drainable porosity was 29.4% and was similar to that of column 1. These results suggest that Ca does slightly affect but not play a significant role in the removal of Cr(VI) by Fe0 in low to moderately hard water.

In column 5, which was fed with Cr(VI) contaminated synthetic groundwater with 6 mM of NaHCO3, the Cr(VI) removal capacity was slightly increased (by 14.6% based on the average values) compared to that of column 1 (Table 1). The drainable porosity was 31.5% and almost the same as that of column 1. The slight increase of Cr(VI) removal capacity in column 5 might be ascribed to the enhanced Fe0 reactivity by bicarbonate. Since the effluent pH in column 5 was below 10 in the column experiment and had an average pH value of 9.72 (Supporting Information Fig. S2a), both bicarbonate (HCO3) and carbonate (CO32−) were the dominated carbonate species with a [HCO3]/[CO32−] ratio of about 4. In Fe0 PRBs, bicarbonate was reported to induce an enhancement of Fe0 reactivity (Gu et al. 1999), which might result from the favored anodic iron corrosion on the iron surface (Castro et al. 1991). Additionally, some hexagonal-shaped morphology was observed in the SEM image of the surface of the reacted Fe0 collected in column 5 (Supporting Information Fig. S3e). This mineral was speculated to be carbonate green rust (Gu et al. 1999; Roh et al. 2000; Lai and Lo 2008). Such carbonate green rust might contribute to Cr(VI) reduction in column 5 because it is an Fe2+-bearing intermediate species of iron corrosion products and was identified as being capable of reducing Cr(VI) in batch kinetics studies (Williams and Scherer 2001; Legrand et al. 2004).

In column 7, which was fed with Cr(VI) contaminated synthetic groundwater containing both CaCl2 and NaHCO3, each gram of Fe0 removed only 2.85 mg of Cr(VI), which was a decrease of 31.4% compared with that of column 1, whereas the drainable porosity in column 7 was 28.5%, which was a slight decrease compared to that of column 1. These results indicated that the combination of Ca and bicarbonate exerted a significant impact on the performance of Fe0, and such impact is more related to the passivation of Fe0 reactivity than the reduction of Fe0 matrix porosity. When compared with those of columns 3 and 5, respectively, noticeable decreases in the Ca concentration and the total alkalinity level (Supporting Information Fig. S2b and c) were recorded in the effluent of column 7. It is suggestive of the formation of calcium carbonate (CaCO3), whose accumulation rate was estimated to be 10.93 mg/PV throughout the column (Supporting Information Table S1). The SEM images revealed cubic-shaped structures on the surface of the reacted Fe0 collected in column 7 (Supporting Information Fig. S3i), which should be calcite. The XRD and FTIR analyses further confirmed the presence of calcite (Supporting Information Figs. S4 and S5). The significant impact of CaCO3 minerals on the performance of Fe0 was widely evidenced (Lo et al. 2006; Lai and Lo 2008). It might be because the CaCO3 precipitates greatly hinder the electron transfer from the Fe0 surfaces to the contaminants and hence passivate the reactivity of Fe0.

3.3 Effects of Humic Acid

3.3.1 Cr(VI) Reduction by Humic Acid

Due to containing some minor reducing functional groups, such as quinones, humic acid could reduce Cr(VI) directly (Tratnyek et al. 2001). A batch experiment was carried out to assess the extent to which Cr(VI) could be directly reduced by humic acid. As shown in Supporting Information Fig. S6, Cr(VI) reduction by humic acid appeared to be highly dependent on solution pH. In acidic and neutral solutions, humic acid can reduce Cr(VI) although at a slow rate (Supporting Information Fig. S6a and b). In the solution with a higher pH (9.4–10.3), which is typical in Fe0 PRBs, little Cr(VI) reduction was observed (Supporting Information Fig. S6c). These results indicate that the reducing capacity of humic acid towards Cr(VI) can be neglected in systems where Fe0 is applied as a predominant reductant to remove Cr(VI).

3.3.2 Effect of Humic Acid in the Absence of Ca: Columns 2 and 6

Figure 3a shows the relative concentrations of Cr(VI) and humic acid, and the concentration of dissolved iron in the effluent of column 2 in the course of the column experiment. Considerable amounts of dissolved iron were released from the column. A similar phenomenon was also observed in column 6 which received Cr(VI) solution with both humic acid and NaHCO3 (Supporting Information Fig. S7a). The significant release of dissolved iron from Fe0 columns fed with humic acid has been reported in a previous study (Dries et al. 2005), though without further explanation. Comparatively, almost no dissolved iron was measured in the columns without the introduction of humic acid (Supporting Information Fig. S2d). Iron corrosion is the essential mechanism for removing Cr(VI) (Powell et al. 1995), in which Fe0 is oxidized to Fe2+ and Fe3+, and Cr(VI) is reduced to Cr(III) with the simultaneous formation of Fe (hydr)oxides as iron corrosion products. Humic acid is well known in having a high binding affinity towards Fe2+ and Fe3+ (Tipping 2002), as well as having a strong tendency to be adsorbed onto iron oxide surfaces (Gu et al. 1994). The adsorbed humic acid could significantly modify the surface electrostatic properties of iron oxide particles (Saito et al. 2004), resulting in the stabilized iron oxide colloids in aqueous phase (von der Kumpulainen et al. 2008). In column 2, some released Fe2+ and/or Fe3+ due to iron corrosion probably bound with humic acid instead of forming Fe (hydr)oxides. Furthermore, some formed Fe (hydr)oxides might be stabilized as fine colloids (<0.45 μm) in solution by the adsorbed humic acid. Hence, a considerable amount of dissolved iron was released from the column, which comprised soluble Fe-humic acid complexes and humic acid stabilized fine Fe (hydr)oxide colloids, resulting in a lower Fe(hydr)oxide accumulation rate than that in column 1 (Supporting Information Table S1). A supplementary experiment was conducted using high-speed centrifugation to separate fine colloids (Supporting Information Supplementary Experiment 1). It was found that in the water samples collected from the effluent of column 2, about two-thirds of the dissolved iron was in the form of fine colloids (<0.45 μm) and the remaining one-third might be in the form of soluble Fe-humic acid complexes. In this column study, the highest dissolved iron concentration was about 3 mg l−1, while in a previous batch study (Liu et al. 2008), the dissolved iron concentration reached 25 mg l−1. It might be ascribed to the much higher Fe0 to solution ratio in the column study, where more solid surfaces were available for the deposition of fine Fe (hydr)oxide colloids.
https://static-content.springer.com/image/art%3A10.1007%2Fs11270-010-0546-2/MediaObjects/11270_2010_546_Fig3_HTML.gif
Fig. 3

a The relative concentrations of Cr(VI) and humic acid, and the concentration of dissolved iron in the effluent column 2, and b the relative concentrations of Cr(VI), humic acid, and Ca, and the concentration of dissolved iron in the effluent column 4 in the course of the column experiments

By forming soluble Fe-humic acid complexes and stabilized fine Fe (hydr)oxide colloids, humic acid could suppress the precipitation of iron corrosion products (Liu et al. 2008) and then diminish the accumulation of secondary minerals on the Fe0 surfaces, as reflected by the higher drainable porosities in columns 2 and 6 than in columns 1, 3, 5, and 7 (Table 1). However, the effect of humic acid in enhancing the Cr(VI) removal capacity was not significant. In column 2, the presence of humic acid slightly increased the Cr(VI) removal capacity by 18.4% compared with that of column 1, whereas in column 6, it was lowered by 7.9%. The underlying reason might be that even though humic acid could suppress the precipitation of iron corrosion product, it may not be able to prevent the passivation on iron surfaces. Once the reactive sites on iron surface were covered by secondary minerals, as evidence in SEM images (Fig. 2c and d), the buildup of iron corrosion products would deteriorate Fe0 reactivity and then decrease Cr(VI) removal capacity, despite the extent of secondary mineral accumulation might be diminished by humic acid. It was further supported by the changes of the relative concentrations of Cr(VI) and humic acid, and the concentration of dissolved iron in the effluent of column 2 (Fig. 3a). As shown in Fig. 3a, during the column experiment, a drop in dissolved iron concentration occurred in the effluent of column 2, and was followed by the detection of Cr(VI), whose concentration increased rapidly thereafter. However, the concentration of humic acid did not show a close correlation to that of Cr(VI) and dissolved iron. Humic acid breakthrough was achieved at about 200 PV, suggesting plenty of humic acid in solution. Since iron corrosion is the source of release of Fe2+ and Fe3+ and is inhibited by the accumulation of iron corrosion products on the Fe0 surfaces, the decrease in the dissolved iron concentration was believed to be due to the passivation of the iron surfaces rather a shortage of the amount of humic acid that was available for binding with Fe2+ and Fe3+ and stabilizing the fine Fe (hydr)oxide colloids. The close correlation of the decrease in the dissolved iron concentration and the increase in the Cr(VI) concentration also suggests the deterioration of Fe0 reactivity by the buildup of iron corrosion products.

The shapes of the humic acid breakthrough curves (Fig. 3a and Supporting Information Fig. S7a) suggest that the humic acid uptake in columns 2 and 6 mainly result from adsorption. The adsorption of humic acid onto iron oxides is believed to involve ligand exchange interactions between the active sites of iron oxide surfaces (Fe–OH) and the deprotonated functional groups (e.g., carboxyls and phenolic hydroxyls) of humic acid (Gu et al. 1994). In the present study, the extent of humic acid adsorption was limited and might be due to the high pH in the columns and the small surface area of the Fe0 used (1.8 m2 g−1); therefore humic acid breakthrough was achieved rapidly after starting the experiments of columns 2 and 6.

3.3.3 Effect of Humic Acid in the Presence of Ca: Columns 4 and 8

By contrast, in columns 4 and 8 where Ca was received, the presence of humic acid resulted in large decreases of the Cr(VI) removal capacity of 24.4% and 42.7%, respectively, compared with column 1. In a previous batch study (Liu et al. 2008), only a slight decrease of the Cr(VI) reduction rates (10–12%) was observed in solutions with co-present Ca and humic acid. This difference suggests that more severe adverse effects could be induced by the co-present Ca and humic acid in a Fe0 matrix closer resembling a real Fe0 PRB. Furthermore, the drainable porosities in columns 4 and 8 were much less than in columns 1, 3, 5, and 7 (Table 1). These results strongly suggest the adverse effects of co-present Ca and humic acid on the performance of Fe0. Figure 3b shows the relative concentrations of Cr(VI), Ca, and humic acid, and the concentration of dissolved iron in the effluent of column 4. Little dissolved iron was released from the column. Unlike in columns 2 and 6, the humic acid concentration showed a close correlation to that of Cr(VI). Almost simultaneously, the rapid increases of Cr(VI) and humic acid concentration took place. A similar phenomenon was also observed in column 8 (Supporting Information Fig. S7b). In columns 4 and 8, humic acid breakthrough was never achieved, suggesting a lot of humic acid was retained in the column. The concentration of Ca did not correlate to those of Cr(VI) and humic acid, which reached a plateau (around C/C0 = 0.7) at about 200 PV.

As aforementioned, in a Ca-free solution, humic acid could bind with Fe2+ and/or Fe3+ and stabilize fine Fe (hydr)oxide colloids, resulting in the detection of a considerable amount of dissolved iron. However, in the presence of Ca, the release of dissolved iron appeared to be significantly affected. Some Ca in solution can be adsorbed onto the fine Fe (hydr)oxide colloids and counteract the effects of humic acid by charge neutralization (Weng et al. 2005). The surface charge of fine Fe (hydr)oxide colloids might be eventually reversed (Duan and Gregory 2003), and hence the aggregation of Fe (hydr)oxide colloids was largely promoted. In addition, Ca can also bind with humic acid and lead to a more compact conformation of humic acid molecules by reducing intramolecular electrostatic repulsions and forming multidentate complexes with neighboring functional groups in a humic acid molecule (Tipping 2002). The Fe (hydr)oxide colloids and humic acid were very likely to co-aggregate, and finally deposit on the surface of Fe0. The SEM images of the surface and the cross-section of the reacted Fe0 collected in column 4 (Fig. 2e and f) clearly revealed a paste-like solid phase, which was believed to be the deposition of aggregated humic acid and with an estimated accumulation rate of 5.46 mg/PV throughout column 4 (Supporting Information Table S1). The increased carbon concentration indicated by EDX analysis also suggested the buildup of humic acid. FTIR analysis further confirmed this point (Supporting Information Fig. S8). In the difference FTIR spectra of the surface coatings of the reacted Fe0 collected in column 4, two new peaks appeared at 2916 and 2848 cm−1, indicating the abundance of aliphatic groups that are not expected to be involved in humic acid adsorption (Gu et al. 1994) and whose prevalent presence should be a result of the deposition of aggregated humic acid. A similar result was also observed in column 8 (Supporting Information Fig. S9).

It should be noted that Ca itself was not able to promote humic acid aggregation to such a great extent that little humic acid was released from column 4 in the first 500 PV of the column experiment (Fig. 3b). As suggested by the result of a supplementary experiment (Supporting Information Supplementary Experiment 2), the enhanced humic acid aggregation should be a synergic process involving both Ca and Fe (hydr)oxide colloids, where the Fe (hydr)oxide colloids are a kind of “coagulant” or “adsorbent” towards humic acid, and Ca acts as an “aid” to promote the co-aggregation Fe (hydr)oxide colloids and humic acid. The simultaneous rapid increases of Cr(VI) and humic acid concentration in columns 4 and 6 (Fig. 3b and Supporting Information Fig. S7b) signified the role of Fe (hydr)oxides. Along with the deposition of Fe (hydr)oxides and humic acid aggregates, iron corrosion was largely diminished and the electron transfer from the Fe0 surfaces to Cr(VI) was greatly inhibited. The increased Cr(VI) concentration suggested the diminishing of iron corrosion from which Fe2+ and Fe3+ were released. As a result, fewer Fe (hydr)oxide colloids were generated from Fe2+ and Fe3+ hydrolysis, and hence, more humic acid began to flow out of the column. Simultaneously, the Cr(VI) removal capacity of Fe0 depleted rapidly.

4 Conclusion

The results of the column experiments presented above show that the influences of individual Ca or bicarbonate on the performance of Fe0 in removing Cr(VI) were not significant, whereas the co-present Ca and bicarbonate significantly decreased the Cr(VI) removal capacity of Fe0 due to the formation of CaCO3 on the Fe0 surfaces. However, the influences of humic acid on the performance of Fe0 in removing Cr(VI) varied significantly, depending on the solutes present in the simulated groundwater. Since hardness and carbonate species are prevalent in groundwater, their influences have been considered in the application of Fe0 PRBs. This study suggests that in NOM-rich environments, such as downstream of a bog or a landfill, the presence of NOM may be a sufficiently important factor influencing the performance of Fe PRBs, and this consideration should be taken into account in the performance prediction and design of Fe0 PRBs.

Acknowledgment

This work was supported by the Hong Kong Research Grants Council under grant HKUST RGC 617006.

Supporting Information

Additional one table, nine figures and information of two supplementary experiments are available online.

Supplementary material

11270_2010_546_MOESM1_ESM.pdf (479 kb)
ESM 1(PDF 479 kb)

Copyright information

© Springer Science+Business Media B.V. 2010