SEARCH

SEARCH BY CITATION

Keywords:

  • Humic acid;
  • Tetracycline;
  • Pyrene;
  • Porous media;
  • Transport

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

The authors observed that humic acid (HA) mediates transport of tetracycline and pyrene in saturated porous media via distinctively different mechanisms. The presence of HA (20–80 mg C/L) in the influent consistently enhances the transport of tetracycline, whereas for pyrene a critical HA concentration exists (about 10 mg C/L), below which transport is inhibited but above which transport is enhanced. The difference in the HA effect stems from the difference in relative sorption affinity to HA and sand between these two compounds. Because sorption of pyrene is driven primarily by hydrophobic effect, pyrene exhibits much stronger sorption to HA than on sand. Accordingly, pyrene in the influent (or mobile phase) is predominantly associated with HA, and its transport is controlled by the partition of HA between mobile phase and sand. For the polar, ionic, and highly hydrophilic tetracycline, sorption is driven mainly by surface complexation and ligand exchange, so tetracycline exhibits relatively strong adsorption on sand, but has much weaker sorption to HA than pyrene does. For tetracycline, the effect of HA on transport is likely the competition of HA for the available adsorption sites on sand. In addition, tetracycline and pyrene exhibit markedly different breakthrough profiles, both in the presence and in the absence of HA; this can be attributed to the greater degree of adsorption nonequilibrium of tetracycline on sand. Environ. Toxicol. Chem. 2012;31:534–541. © 2011 SETAC


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Antibiotics such as tetracyclines and sulfonamides are produced and used in large quantities worldwide 1. These chemicals are released into soil, aquatic, and groundwater environments from many sources, such as livestock production and discharges from hospitals and pharmaceutical plants 1, 2. The environmental fate and transport of these emerging contaminants have received much attention 3, 4. Nonetheless, although numerous studies have been conducted to understand the fate processes of antibiotics in soil 5 and aquatic environments 6, 7, to date very little information is available on the transport of antibiotics in groundwater environments 1.

Groundwater typically contains a few to a few hundred milligrams per liter of dissolved organic matter (DOM) 8. Many studies have shown that DOM can significantly affect the transport of contaminants in groundwater 8–11. For example, several studies have shown that DOM can serve as a contaminant carrier and therefore enhances the transport of both hydrophobic organic compounds and metals in groundwater 8, 10–13. Several other studies have indicated that DOM can also inhibit the transport of hydrophobic organic contaminants in the case of significant retention of DOM within porous media 9, 14, 15. The significance of DOM's effect on contaminant transport depends not only on the abundance and mobility of DOM in groundwater but also on the nature of contaminant–DOM interaction and contaminant–aquifer material interaction 16, 17. For apolar, nonionic, hydrophobic organic contaminants, sorption to both DOM and aquifer material is controlled primarily by hydrophobic effect 18. Nonetheless, for polar, ionic, and highly hydrophilic organic contaminants such as tetracyclines, significant nonhydrophobic interactions (for example, cation exchange, ligand exchange, surface complexation, and hydrogen bonding) can be predominant 5. For example, the sorption affinities, and thus phase distribution, of tetracycline are not only strongly dependent on the physicochemical properties (e.g., aromaticity, polarity, abundance, and types of functional groups) of DOM and soil organic matter 19, but also complicated by its specific interactions with soil clay minerals 20, 21. Thus far, little is known about the effects of DOM on the transport of antibiotics in subsurface environments. However, it is possible that even the same type of DOM can have effects on the transport of antibiotics markedly different from effects on the transport of nonionic, hydrophobic organic contaminants.

The objective of the present study was to understand whether DOM affects the transport of antibiotics in saturated porous media differently from the way it affects the transport of apolar, nonionic, hydrophobic organic contaminants. Tetracycline and pyrene were used as the model contaminants; the two chemicals differ significantly in hydrophobicity, solubility, polarity, and functional groups. A soil-derived humic acid (HA) was used as the model DOM. The transport characteristics of tetracycline and pyrene through saturated porous media in the presence and absence of HA were studied with column tests and fitted with contaminant transport models. Batch sorption experiments were conducted to understand better the mechanisms controlling the effects of HA on contaminant transport.

MATERIALS AND METHODS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Materials and chemicals

The HA was isolated from a surface soil (0–20 cm) using a method described in the literature 22. The detailed procedures are described in the Supplemental Data. Concentrations of HA are expressed as milligrams of carbon per liter (mg C/L) in the present study. Elemental analysis of freeze-dried HA was performed using an Elementar Vario Micro Cube (Elementar Analysensysteme). Fourier transform infrared spectroscopy analysis was performed using a Bruker Tensor 27 Fourier transform infrared spectrometer (Bruker Optics) with an HA-KBr ratio of 1:100 (w/w). Solid-state 13C nuclear magnetic resonance spectroscopy analysis was performed using a Bruker Avance III 400 (Bruker BioSpin) at 100 MHz with 4-mm rotors in a double-resonance probe head. The relaxation time and acquisition time were 2.5 s and 0.034 s, respectively.

The porous medium was a fine-to-medium-grain sand near the surface collected from the coast of Tianjin, China. The sand has an average particle size of 0.255 mm, a fractional organic carbon (fOC) value of 0.013% (determined with the Walkley–Black method), and a cation-exchange capacity of 13.9 mmol/kg. The mineral analysis of the sand was performed using an AxiosmAX X-ray fluorescence spectrometer (PANalytical).

Tetracycline (purity ≥99%) and pyrene (purity ≥98%) were obtained from Sigma-Aldrich. Selected physiochemical properties of the two compounds are given in Supplemental Data, Table S1.

Column transport experiments

The column experiments were conducted using borosilicate glass columns (Omnifit Bio-Chem Valve). The length of the column was 10 cm, and the inner diameter was 1.0 cm. The column had one adjustable end piece, and a 10-µm stainless steel screen (Valco Instruments) in either side. A KD model 210 syringe pump (KD Scientific) equipped with two 50-ml glass syringes with steel plungers (Hamilton) was used to flush solutions into the columns. The whole system was connected using 1/16-inch PTFE tubing, and the column was wrapped with aluminum foil. The schematic illustration of the experimental apparatus is shown in Supplemental Data, Figure S1.

The experimental protocols of the column tests are summarized in Table 1. For each column, approximately 7 g sand was dry packed into the column and flushed with deionized water backward at a flow rate of 1.0 ml/h for several days to saturate the column. The length of the packed sand was approximately 5 cm. The breakthrough curves of NaBr, used as a conservative tracer, were used to calibrate the dead volume (∼0.5 ml) and pore volume of the columns.

Table 1. Experimental protocols of column tests and fitted parameters of two-site nonequilibrium transport model
Experiment No.ContaminantColumn propertiesInfluent propertiesFitted parameters of two-site nonequilibrium transport modela
ρb (g/cm3)θ (−)v (m/d)pHTotal humic acid concn. (mg C/L)Total cont. concn. (mg/L)Sorbed concn.b (%)Dc (m2/d)R (−)β (−)ω (−)Kd (L/kg)f (−)α (1/d)r2
  • a

    Standard errors associated with the fitted parameters are given in Supplemental Data, Table S5.

  • b

    Calculated using the Kc-HA (L/kg C) values of 8.29 × 104 for pyrene and 939 for tetracycline (Fig. 4): sorbed cont. = Kc-HACHA/(1 + Kc-HACHA) × 100.

  • c

    Obtained by fitting the breakthrough data of NaBr, a conservative tracer (Supplemental Data, Fig. S4).

1Pyrene1.760.3482.376.900.0630.00.51528.50.5811.165.440.5664.380.996
2Pyrene1.760.3482.377.1200.06362.40.51533.50.5661.076.430.5533.290.998
3Pyrene1.760.3482.377.0500.06380.60.51519.00.4621.153.540.4325.090.986
4Pyrene1.760.3482.377.2800.06386.90.51516.80.4911.293.130.4596.770.987
5Tetracycline1.700.3712.226.8010.00.00.48367.40.1962.3014.50.1851.910.996
6Tetracycline1.700.3712.227.02010.01.840.48354.90.2542.0311.80.2412.220.997
7Tetracycline1.700.3712.227.05010.04.480.48334.70.2352.307.350.2143.890.997
8Tetracycline1.700.3712.226.98010.06.990.48334.40.1562.417.290.1333.720.998

The influent solution of a typical column experiment was prepared by adding a stock solution of tetracycline (in deionized water) or pyrene (in methanol) to an electrolyte solution (0.01 M NaCl and 0.01 M NaN3), containing a certain amount of HA (0, 20, 50, or 80 mg C/L), to give an initial concentration of 0.063 mg/L for pyrene or 10 mg/L for tetracycline. (The mass fraction of methanol was kept below a volumetric ratio of 0.1% to minimize cosolvent effects.) The mixture was stirred to allow sorption equilibrium of the contaminant to HA in the influent. While the influent solution was being prepared, the contaminant-free HA electrolyte solution—with the same HA concentration as that of the influent—was flushed continually into the column to equilibrate the sand with HA; only the background electrolyte was used for the experiments in which the influent contained no HA. Then, the prepared influent was flushed into the column. The effluent was collected at predetermined time intervals to measure contaminant concentration and pH. The tetracycline experiments and the pyrene experiments were run separately. In all experiments, the effluent pH was essentially the same as the influent pH.

Batch sorption experiments

Sorption isotherms of tetracycline and pyrene to sand were obtained using a previously developed method 23. First, a series of 40-ml aluminum foil-wrapped amber glass vials each containing 5 g sand and 40-ml electrolyte solution (0.01 M NaCl and 0.01 M NaN3, pH 7.0 ± 0.2) was prepared. For the tetracycline experiments, the electrolyte solution was purged with N2 for 6 h before being used. Different amounts of tetracycline or pyrene stock solution were then added to the vials. The vials were filled with the electrolyte solution immediately to leave minimal headspace and then tumbled end-over-end at 1 rpm for 3 d (tetracycline) or 7 d (pyrene). Afterward, the vials were centrifuged at 3,000 rpm for 30 min, and the supernatant was withdrawn to analyze the concentrations of tetracycline or pyrene. To account for possible solute loss from processes other than sorption to sand, calibration curves were obtained separately from controls receiving the same treatment as the sorption samples but without sand. Calibration curves included at least 14 standards over the tested concentration ranges. Based on the obtained calibration curves, the sorbed mass was calculated by subtracting mass in the aqueous phase from the total mass spiked. The sorption isotherms of tetracycline and pyrene to sand in the presence of HA were obtained using the same procedures, but the electrolyte solution also contained different concentrations of HA (20, 50, and 80 mg C/L, respectively).

The sorption isotherm of tetracycline to HA was obtained using the method of Gu et al. 24. First, a series of dialysis bags (1,000 molecular weight cutoff) each containing 5 ml of 320 mg C/L HA solution was placed in 40-ml amber glass vials containing approximately 40-ml background electrolyte solution. Then, a certain amount of tetracycline stock solution was spiked into the electrolyte solution in each vial. Afterward, the vials were filled with the electrolyte solution, sealed, and tumbled for 3 d to allow tetracycline molecules to pass through the dialysis membrane and to reach sorption equilibrium. Then, tetracycline concentrations inside and outside of dialysis bags were measured. The concentration inside the bag represents the bulk concentration of both freely dissolved and HA-associated tetracycline, and the concentration outside the bag represents the concentration of freely dissolved tetracycline. The concentration of HA-associated tetracycline was calculated based on mass balance (see Supplemental Data). The sorption isotherm of pyrene to HA was obtained using a slightly different approach 25. For each experiment, approximately 5 ml electrolyte solution was sealed in a dialysis bag and was placed in a 200-ml amber glass bottle containing approximately 200 ml of 10 mg C/L HA in the background electrolyte solution. Then, a pyrene stock solution was spiked in the solution outside of the dialysis bag. The bottle was sealed and tumbled for 7 d. Then, pyrene concentrations inside and outside of the dialysis bag were measured. The concentration inside the bag represents the concentration of freely dissolved pyrene, and the concentration outside the bag represents the bulk concentration of both freely dissolved and HA-associated pyrene. The sorption coefficients of pyrene to HA were also obtained using the solid-phase dosing method of ter Laak et al. 26. The detailed procedures and calculations are given in the Supplemental Data.

To obtain the adsorption isotherm of HA on sand, 5 g sand was added to a series of 40-ml amber glass vials each containing different concentrations of HA. The vials were tumbled end-over-end at 1 rpm for 3 d. Then, the vials were centrifuged, and HA concentrations in the supernatant were measured. The concentrations of HA in the sorbed phase were calculated from the total mass spiked and the measured aqueous phase concentrations based on mass balance.

Each sorption data point was run in duplicate. The blank control samples of tetracycline and pyrene showed no degradation during sorption experiments, and pH remained constant during all of the experiments.

Analytical methods

The concentrations of pyrene and tetracycline were determined using a Waters high-performance liquid chromatography system equipped with a symmetry reversed-phase C18 column (4.6 × 150 mm; Waters). Pyrene was detected with a Waters 2475 fluorescence detector at an excitation wavelength of 325 nm and an emission wavelength of 390 nm; the mobile phase was methyl cyanide–deionized water (85:15, v:v; 1.0 ml/min). Tetracycline was detected with a Waters 2487 UV/visible detector at a wavelength of 360 nm; the mobile phase was methyl cyanide–2% oxalic acid in water (15:85, v:v; 1.0 ml/min). The organic carbon concentration of HA was determined by using a TOC-V CPH analyzer (Shimadzu).

Contaminant transport models

Two one-dimensional transport models were used to fit the data of the column experiments. The first model is the advection–dispersion equation assuming the sorption of contaminant is at equilibrium 27

  • equation image(1)
  • equation image(2)

where R (unitless) is the retardation factor resulting from sorption, C (mg/L) is the solution-phase concentration, ρb (g/cm3) and θ (unitless) are the bulk density and porosity of the porous medium, Kd (L/kg) is the contaminant sorption coefficient, D (m2/d) is dispersion coefficient, v (m/d) is linear velocity, t (d) is time, and x (m) is distance.

The second transport model is based on the assumption that two types of sorption sites exist, an equilibrium site and a kinetic site 28

  • equation image(3)
  • equation image(4)

where Sk (mg/kg) is the sorbed-phase concentration of a contaminant in the kinetic site, α (1/d) is the first-order rate coefficient associated with the kinetic site, and f (unitless) is the fraction of sites for which sorption is at equilibrium. Note that a constant Kd value is assumed for the two sites. The second model can be reduced to a dimensionless form as 28, 29

  • equation image(5)
  • equation image(6)

where

  • equation image(7)
  • equation image(8)

and where the subscripts 1 and 2 refer to the equilibrium and kinetic sites, respectively; T and X are dimensionless time and distance, respectively; P is the Peclet number (P = vL/D); β is fraction of instantaneous retardation; ω is the Damkohler number (i.e., ratio of hydrodynamic residence time to characteristic time for sorption); and L (m) is the length of the porous medium.

The values of D were obtained by fitting the breakthrough data of the conservative tracer 27 using the CXTFIT 2.1 code 30, and the values of R, β, and ω were obtained by fitting the transport data of the test contaminants. The values of Kd, f, and α were calculated using Equations 2, 7, and 8.

RESULTS AND DISCUSSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Characterization of HA and porous media

The elemental analysis results for HA (Supplemental Data, Table S2) show that HA has an O/C ratio of 0.58 and an (O + N)/C ratio of 0.65. The Fourier transform infrared spectra of HA (Supplemental Data, Fig. S2) confirm the existence of carboxylic, phenolic, and alcohol groups. The 13C NMR spectra of HA (Supplemental Data, Fig. S3 and Table S3) also indicate that HA has high contents of polar functional groups, approximately 11.3% carboxylic C (chemical shift 163–190 ppm), 9.0% phenolic C (145–163 ppm), 21% O-alkyl (61–109 ppm), and 9.2% N-alkyl/methoxyl (50–61 ppm). The mineral analysis results for the sand (Supplemental Data, Table S4) indicate that the major mineral components of the material are SiO2 (80% mass fraction) and Al2O3 (8.6%); the iron oxide content of the sand is approximately 0.82%. The breakthrough curves of NaBr, the conservative tracer, are shown in Supplemental Data, Figure S4, and the fitted D values are listed in Table 1.

Effects of HA on contaminant transport

The breakthrough curves of pyrene and tetracycline in the presence and absence of HA are compared in Figure 1. Note that in the presence of HA in the mobile phase, the solution-phase concentration of tetracycline or pyrene, C, is a bulk concentration for both freely dissolved contaminant and contaminant sorbed to dissolved HA. For both pyrene and tetracycline, the breakthrough data can be well described with the two-site nonequilibrium transport model, and the fitted parameters are summarized in Table 1. The transport data for neither contaminant can be described with the equilibrium transport model (Eqn. 1; see Supplemental Data, Fig. S5).

thumbnail image

Figure 1. Breakthrough curves of pyrene (a) and tetracycline (b) in the presence and absence of humic acid (HA). Column and influent properties are summarized in Table 1. The dashed lines were plotted by curve-fitting experimental data with the two-site nonequilibrium transport model (Eqns. 5 and 6). PV = pore volumes.

Download figure to PowerPoint

In the absence of HA, breakthrough of pyrene occurred at 13 pore volumes. The percentage of breakthrough—C/C0 (where C0 is the total contaminant concentration in the influent; Table 1)—increased quickly and reached more than 80% after 40 pore volumes. In the presence of 20 mg C/L HA in the influent, the transport of pyrene was inhibited: the R value increased by 18% (Table 1). Nonetheless, in the presence of 50 and 80 mg C/L HA, the transport of pyrene was facilitated, and the R value decreased by more than 33 and 41%, respectively.

The breakthrough curves for tetracycline exhibit shapes markedly different from those for pyrene. In the absence of HA, tetracycline breakthrough occurred after 10 pore volumes. However, as the flow continued, the C/C0 ratio of tetracycline increased much more slowly compared with that of pyrene, and after 40 pore volumes the C/C0 ratio was less than 40%. This indicates that, at a given pore volume, a larger fraction of tetracycline than pyrene was retained by the porous medium. More important, comparing the breakthrough profiles between tetracycline and pyrene indicates that the transport properties of tetracycline deviate much more significantly from the equilibrium-based transport model (Eqn. 1) 31, 32. This can be understood further with the fitted parameters in Table 1, in that for tetracycline the fraction of the nonequilibrium-sorption site (that is, 1−f) is much larger, and the kinetic constant associated with the nonequilibrium site (α) is smaller. (The slower kinetics of tetracycline can possibly be linked to its greater polarity 33.) Figure 1 also shows that HA has a distinctively different effect on the transport of tetracycline. Contrary to the case for pyrene, the effect of HA on the transport of tetracycline was consistent: with the increase of HA concentration, the breakthrough curve shifts consistently to the left. Thus, for tetracycline, transport is always enhanced by the presence of HA.

Effects of HA on contaminant sorption

A comparison of the fitted R values in Table 1 indicates that low HA concentrations in the influent enhance sorption of pyrene to sand, which leads to inhibited transport, but high HA concentrations inhibit sorption of pyrene, which leads to enhanced transport; nonetheless, even low HA concentrations in the influent inhibit adsorption of tetracycline on sand, so tetracycline transport is always enhanced in the presence of HA. These assumptions were verified with the batch sorption data of pyrene and tetracycline to sand in the presence and absence of HA (Fig. 2). Note that, in this set of experiments, pyrene and tetracycline molecules are distributed among four phases (refer to the schematic illustration in Fig. 3), including molecules freely dissolved in solution (phase 1), molecules sorbed to dissolved HA (phase 2), molecules sorbed to sand (phase 3), and molecules sorbed to HA that is adsorbed on sand (phase 4). In Figure 2, the aqueous-phase concentration (C) is the bulk contaminant concentration of phases 1 and 2, and the sorbed-phase concentration (q) is the bulk concentration of phases 3 and 4.

thumbnail image

Figure 2. Sorption isotherms of pyrene (a) and tetracycline (b) to sand in the presence and absence of humic acid (HA). The HA concentrations (20, 50, and 80 mg C/L) are total HA concentration in the system. Solid lines were plotted by fitting the data with linear sorption isotherm.

Download figure to PowerPoint

thumbnail image

Figure 3. Schematic illustration of contaminant mass distribution among four phases (each indicated with Roman numbers I, II, III, or IV) in a closed system containing sand, aqueous solution, contaminant, and humic acid.

Download figure to PowerPoint

Figure 2 clearly shows that, in the presence of 20 mg C/L HA, sorption of pyrene to sand was considerably enhanced compared with the sorption in the absence of HA. In the presence of 50 or 80 mg C/L HA, sorption of pyrene was significantly inhibited. Nonetheless, HA at all three concentrations inhibited the adsorption of tetracycline on sand. In addition, the adsorption isotherms of tetracycline in the presence of 50 and 80 mg C/L HA nearly overlap, indicating that, once the HA concentration is above a certain level, its adsorption-inhibition effect stops increasing further. In Table 2, the Kd values obtained in the batch sorption experiments are compared with the Kd values calculated from the R values (Table 1) that were obtained by fitting the transport data with the two-site model (Eqns. 5–8). A sensitive analysis of the model parameter R was performed, and the results are shown in Supplemental Data, Fig. S6. In general, the two groups of Kd values agree well for both pyrene and tetracycline.

Table 2. Comparison of experimentally obtained and estimated Kd values of pyrene and tetracycline to sand in the presence and absence of humic acid (HA)
CHA_total (mg C/L)Kd (L/kg)a
PyreneTetracycline
ColumnbBatchcEstimateddColumnbBatchcEstimatedd
  • a

    Values in parentheses represent standard errors.

  • b

    Calculated from the fitted R value using Equation 2; R was obtained by fitting the transport data with the two-site model (Eqns. 5–8).

  • c

    Measured in batch sorption experiments.

  • d

    Calculated Kd_apparent values using Equation 13.

05.44 (0.05)4.81 (0.08)4.8114.5 (0.1)14.9 (0.4)14.9
206.43 (0.06)5.93 (0.08)5.5811.8 (0.1)12.5 (0.4)14.8
503.54 (0.08)3.26 (0.04)3.317.35 (0.05)6.85 (0.14)14.5
803.13 (0.07)2.73 (0.03)2.217.29 (0.05)6.72 (0.18)14.1

Mechanistic aspects

The most striking observation in the present study is that for pyrene a critical HA concentration appears to exist, below which HA enhances sorption of pyrene to sand but above which HA inhibits sorption; for tetracycline, however, HA always inhibits adsorption. This interesting difference in the effect of HA between pyrene and tetracycline can be understood by examining contaminant mass distribution among the four phases mentioned above. The following equations can be derived based on mass balance (see Supplemental Data for detailed derivation)

  • equation image(9)
  • equation image(10)
  • equation image(11)
  • equation image(12)

where f1 through f4 are mass fraction of a contaminant in each of the four phases; ms/Vw (kg/L) is the sand-to-water ratio; Kc-HA (L/kg C) and Kc-sand (L/kg) are the sorption coefficients of a contaminant to HA and to sand, respectively; CHA (kg C/L) is the concentration of HA in the solution; and qHA (kg C/kg) is the concentration of HA adsorbed on sand. In addition, an apparent Kd value, Kd_apparent, can be calculated as (see Supplemental Data)

  • equation image(13)

where Kd_apparent has the same physical meaning as the slopes of the sorption isotherms in Figure 2 (i.e., it is the ratio of q, the bulk concentration of phases 3 and 4, to C, the bulk contaminant concentration of phases 1 and 2). The values of f1 through f4 and Kd_apparent can be calculated once the values of Kc-HA, Kc-sand, CHA, and qHA have been obtained (these values were obtained experimentally in the present study). Note that two underlying assumptions are that the organic C originally in the sand and the HA that adsorbs to the sand function independently and that sorption affinity of pyrene and tetracycline to the adsorbed HA is the same as that to the aqueous-phase HA (that is, Kc-HA).

Figure 4a shows the sorption isotherm of pyrene to HA. The sorption data follow reasonably the linear sorption isotherm q = Kc-HAC, and a Kc-HA value of 104.92 L/kg C can be obtained. This value is consistent with the literature-reported values 26 and with the sorption coefficients determined using the solid-phase dosing method (Supplemental Data, Fig. S7). A Kc-sand value of 4.81 can be calculated for pyrene using the sorption data in the absence of HA in Figure 2a and can be converted to a KOC value of 104.57 L/kg. Even though sorption of pyrene to both HA and sand is controlled primarily by hydrophobic partitioning to natural organic matter, it is possible that sorption to HA is enhanced by π–π electron donor–acceptor interactions resulting from the aromatic nature of HA 34.

thumbnail image

Figure 4. Sorption isotherms of pyrene (a) and tetracycline (b) to humic acid.

Download figure to PowerPoint

Figure 4b shows the sorption isotherm of tetracycline to HA, and an average Kc-HA value of 939 L/kg C can be obtained. This value is comparable with the literature values 24. A Kc-sand value of 14.9 L/kg was obtained for tetracycline with the sorption data in Figure 2b. These values indicate that tetracycline exhibits much weaker (approximately two orders of magnitude) sorption to HA than pyrene does, but tetracycline adsorbs more strongly on sand than pyrene does. As an amphoteric compound, tetracycline has three ionizable functional groups (Supplemental Data, Fig. S8), each with a specific pKa value 5. Within the test pH, tetracycline exists mainly as a zwitterionic species and can interact strongly with the deprotonated sites of HA (mainly carboxylic groups) 19, 24, 35. Furthermore, because of the high content of polar functional groups in HA (phenolic, carboxylic, and alcoholic groups; Supplemental Data, Table S3), strong H-bonding among the hydroxyl, ketone, and amino groups in tetracycline and the respective functional groups in HA likely is an important sorption mechanism 19, 24, 35. Thus, even though KOW (n-octanol–water partition coefficient) of tetracycline is more than six orders of magnitude lower than that of pyrene, it still exhibits moderate sorption affinity to HA. Similarly, because the interaction of tetracycline with the quartz-like sand is likely via surface complexation and ligand exchange 36, 37 rather than hydrophobic partitioning, tetracycline exhibits stronger adsorption to the extremely low-fOC sand than pyrene does.

Figure 5 shows the adsorption isotherm of HA on sand. The adsorption data can be well described with the Langmuir adsorption isotherm

  • equation image(14)

where qmax (kg C/kg) is the maximum monolayer adsorption capacity of HA on sand and b (L/kg C) is the Langmuir adsorption affinity. This is consistent with the literature 38, in that soil minerals provide surfaces on which amphiphilic humic moieties form a membrane-like coating. The values of b and qmax, obtained by fitting the adsorption data, are 1.1 × 105 L/kg C and 1.3 × 10−4 kg C/kg, respectively. Note that other types of DOM might have different b and qmax values but likely would follow the Langmuir-type adsorption isotherm too. The values of CHA and qHA at a given total HA concentration in the system (CHA_total) can then be calculated based on mass balance as well as Equation 14 (see Supplemental Data).

thumbnail image

Figure 5. Adsorption isotherm of humic acid (HA) on sand. Solid line was plotted by curve-fitting adsorption data with Langmuir model (Eqn. 14).

Download figure to PowerPoint

In Figure 6, the changes of f1 to f4 with CHA_total are shown for pyrene and tetracycline. An interesting observation is that, for pyrene, the value of f4 first increases with CHA_total and then decreases, but no such trend is observed for tetracycline. The Kd_apparent of pyrene shows a very similar pattern (Fig. 6a); it increases with CHA_total and peaks when CHA_total reaches approximately 10 mg C/L, then starts to decrease with the further increase of CHA_total. However, for tetracycline, Kd_apparent is highest when CHA_total is zero and continues to decrease with the increase of CHA_total (Fig. 6b). These Kd_apparentCHA_total correlations provide a good explanation for the observed HA effects on transport (Fig. 1) and on sorption (Fig. 2) that differ between pyrene and tetracycline.

thumbnail image

Figure 6. Mass fraction of pyrene (a) and tetracycline (b) among four phases in response to different total humic acid (HA) concentration in the system (CHA_total). f1 through f4 are defined in Equations 9 to 12, and Kd_apparent is defined in Equation 13.

Download figure to PowerPoint

The Kd_apparent peak or the critical HA concentration for pyrene exists because of a combined effect: the large Kc-HA to Kc-sand ratio and the Langmuir-type adsorption of HA on sand. When HA is present at low concentrations, corresponding to the linear part of the Langmuir isotherm of HA on sand, a large fraction of HA and hence a significant fraction of pyrene (because pyrene sorbs predominantly to HA) are associated with sand, resulting in a higher Kd_apparent compared with that in the absence of HA. Once CHA_total reaches the critical concentration, the adsorbed concentration of HA on sand (qHA) does not increase appreciably with the further increase of CHA_total; that is, qHA approaches the plateau part of the Langmuir isotherm. Accordingly, the fraction of dissolved HA increases, along with sorbed pyrene, and Kd_apparent starts to decrease.

In Table 2, the Kd_apparent values of pyrene and tetracycline estimated using Equation 13 are compared with the values observed in the batch sorption and transport experiments. For pyrene, the estimated values agree reasonably with the experimentally observed values. Nonetheless, the estimated Kd_apparent values for tetracycline deviate increasingly from the observed values with the increase of CHA_total: according to Figure 6, Kd_apparent should change little within the test concentration range of HA, because the mass fractions of tetracycline associated with HA (f2 and f4) should be insignificant compared with the fraction in the solution (f1) and the fraction adsorbed on sand (f3). This is apparently contradictory to the strong effects of HA on the transport and adsorption of tetracycline (Figs. 1 and 2). A possible explanation is the competitive adsorption between HA and tetracycline for the available adsorption sites (most likely silicon or aluminum hydroxyls) on sand 39, 40. As the concentration of HA increases, an increasing fraction of available adsorption site on sand surface is covered by HA, resulting in inhibited adsorption of tetracycline on sand. The fact that the extent of adsorption-inhibition is similar in the presence of 50 and 80 mg C/L HA (Fig. 2) is consistent with the competitive adsorption theory: because adsorption of HA on sand follows the Langmuir isotherm, once qHA approaches the maximum monolayer adsorption capacity, adsorption of HA, and subsequently its competition effect, stops increasing further. Because competitive adsorption is not accounted for in Equations 9 through 13, these equations cannot accurately quantify the effect of HA on tetracycline adsorption on sand, whereas, for pyrene, competitive sorption of HA is likely negligible, because HA and pyrene sorb to different sites on sand, and Eqns. 9–13 work well.

The findings in the present study indicate that the same type of DOM might have effects on the subsurface transport of ionic, polar, and hydrophilic organic contaminants, such as tetracycline antibiotics, considerably different from those on the transport of nonionic, apolar, and highly hydrophobic organic contaminants, such as high-molecular-weight polycyclic aromatic hydrocarbons. The specific effect of DOM on transport depends on the nature of contaminant–DOM, contaminant–porous medium and DOM–porous medium interactions. In general, DOM mediates the transport of nonionic, apolar, highly hydrophobic contaminants primarily by affecting the partition of contaminants between the mobile phase and the porous medium, whereas the effect of DOM on the transport of ionic, polar, and highly hydrophilic contaminants might also involve the competition of DOM for the available adsorption sites on the porous medium, at least when DOM is present at relatively high concentrations. Furthermore, the DOM-mediated transport of emerging contaminants such as antibiotics cannot be accurately quantified with the conventional models and must be better understood.

SUPPLEMENTAL DATA

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Figure S1. Column apparatus.

Figures S2, S3. Fourier transform infrared and solid-state 13C NMR spectra of HA.

Figure S4. Breakthrough curves of tracer.

Figure S5. Breakthrough curves of tetracycline and pyrene fitted with Equation 1.

Figure S6. Sensitivity analysis results of the parameter R of the two-site nonequilibrium model.

Figure S7. Sorption coefficients of pyrene to HA determined using the solid-phase dosing method.

Figure S8. Ionizable groups of tetracycline.

Table S1. Selected physiochemical properties of tetracycline and pyrene.

Tables S2, S3. Elemental analysis and solid-state 13C NMR results of HA.

Table S4. Mineral analysis results of sand.

Table S5. Standard errors associated with the fitted parameters of the two-site nonequilibrium transport model (19.5 KB DOC).

Acknowledgements

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

This project was supported by the National Natural Science Foundation of China (grants 20637030, 20977050, and 21177063), the Tianjin Municipal Science and Technology Commission (grant 10SYSYJC27200), and the China–U.S. Center for Environmental Remediation and Sustainable Development. We thank J. Xiang of Institute of Chemistry, Chinese Academy of Sciences, for his help with solid-state 13C NMR analysis.

REFERENCES

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information
  • 1
    Sarmah AK, Meyer MT, Boxall ABA. 2006. A global perspective on the use, sales, exposure pathways, occurrence, fate and effects of veterinary antibiotics (VAs) in the environment. Chemosphere 65: 725759.
  • 2
    Segura PA, Francois M, Gagnon C, Sauve S. 2009. Review of the occurrence of anti-infectives in contaminated wastewaters and natural and drinking waters. Environ Health Perspect 117: 675684.
  • 3
    Boxall ABA, Kolpin DW, Halling-Sorensen B, Tolls J. 2003. Are veterinary medicines causing environmental risks? Environ Sci Technol 37: 286A294A.
  • 4
    Pan B, Ning P, Xing B. 2009. Part V—Sorption of pharmaceuticals and personal care products. Environ Sci Pollut Res 16: 106116.
  • 5
    Tolls J. 2001. Sorption of veterinary pharmaceuticals in soils: A review. Environ Sci Technol 35: 33973406.
  • 6
    Kuemmerer K. 2009. Antibiotics in the aquatic environment–A review–Part I. Chemosphere 75: 417434.
  • 7
    Kuemmerer K. 2009. Antibiotics in the aquatic environment–A review–Part II. Chemosphere 75: 435441.
  • 8
    Kan AT, Tomson MB. 1990. Groundwater transport of hydrophobic organic compounds in the presence of dissolved organic matter. Environ Toxicol Chem 9: 253263.
  • 9
    Totsche KU, Koegel-Knabner I. 2004. Mobile organic sorbent affected contaminant Transport in soil: numerical case studies for enhanced and reduced mobility. Vadose Zone J 3: 352367.
  • 10
    Ryan JN, Elimelech M. 1996. Colloid mobilization and transport in groundwater. Colloids Surf A 107: 156.
  • 11
    Sen TK, Khilar KC. 2006. Review on subsurface colloids and colloid-associated contaminant transport in saturated porous media. Adv Colloid Interface Sci 119: 7196.
  • 12
    Dunnivant FM, Jardine PM, Taylor DL, McCarthy JF. 1992. Cotransport of cadmium and hexachlorobiphenyl by dissolved organic carbon through columns containing aquifer material. Environ Sci Technol 26: 360368.
  • 13
    Roy SB, Dzombak DA. 1997. Chemical factors influencing colloid-facilitated transport of contaminants in porous media. Environ Sci Technol 31: 656664.
  • 14
    Koegel-Knabner I, Totsche KU. 1998. Influence of dissolved and colloidal phase humic substances on the transport of hydrophobic organic contaminants in soil. Phys Chem Earth 23: 179185.
  • 15
    Totsche KU, Koegel-Knabner I, Danzer J. 1997. Dissolved organic matter-enhanced retention of polycyclic aromatic hydrocarbons in soil miscible displacement experiments. J Environ Qual 26: 10901100.
  • 16
    Kretzschmar R, Borkovec M, Grolimund D, Elimelech M. 1999. Mobile subsurface colloids and their role in contaminant transport. Adv Agron 66: 121193.
  • 17
    Hofmann T, von der Kammer F. 2009. Estimating the relevance of engineered carbonaceous nanoparticle facilitated transport of hydrophobic organic contaminants in porous media. Environ Pollut 157: 11171126.
  • 18
    Schwarzenbach R, Gschwend P, Imboden D. 2003. Environmental Organic Chemistry, 2nd ed. John Wiley & Sons, Inc., Hoboken, NJ, USA.
  • 19
    Sun H, Shi X, Mao J, Zhu D. 2010. Tetracycline sorption to coal and soil humic acids: An examination of humic structural heterogeneity. Environ Toxicol Chem 29: 19341942.
  • 20
    Carrasquillo AJ, Bruland GL, Mackay AA, Vasudevan D. 2008. Sorption of ciprofloxacin and oxytetracycline zwitterions to soils and soil minerals: Influence of compound structure. Environ Sci Technol 42: 76347642.
  • 21
    Figueroa RA, Leonard A, Mackay AA. 2004. Modeling tetracycline antibiotic sorption to clays. Environ Sci Technol 38: 476483.
  • 22
    Swift RS. 1996. Organic matter characterization. In Methods of Soil Analysis: Chemical Methods. Soil Science Society of America, Madison, WI, USA, pp 10111069.
  • 23
    Zhang D, Zhu D, Chen W. 2009. Sorption of nitroaromatics to soils: Comparison of the importance of soil organic matter versus clay. Environ Toxicol Chem 28: 14471454.
  • 24
    Gu C, Karthikeyan KG, Sibley SD, Pedersen JA. 2007. Complexation of the antibiotic tetracycline with humic acid. Chemosphere 66: 14941501.
  • 25
    McCarthy JF, Jlmenez BD. 1985. Interactions between polycyclic aromatic hydrocarbons and dissolved humic material: Binding and dissociation. Environ Sci Technol 19: 10721076.
  • 26
    ter Laak TL, Durjava M, Struijs J, Hermens JLM. 2005. Solid phase dosing and sampling technique to determine partition coefficients of hydrophobic chemicals in complex matrixes. Environ Sci Technol 39: 37363742.
  • 27
    Fetter CW. 1998. Contaminant Hydrogeology, 2nd ed. Prentice Hall, Upper Saddle River, NJ, USA.
  • 28
    van Genuchten MT, Wagenet RJ. 1989. Two-site/two-region models for pesticide transport and degradation: theoretical development and analytical solutions. Soil Sci Soc Am J 53: 13031310.
  • 29
    Nkedi-Kizza P, Biggar JW, Selim HM, van Genuchten MT, Wierenga PJ, Davidson JM, Nielsen DR. 1984. On the equivalence of two conceptual models for describing ion exchange during transport through an aggregated oxisol. Water Resour Res 20: 11231130.
  • 30
    N. Toride FJL, Van Genuchten MT. 1999. The CXTFIT Code for Estimating Transport Parameters From Laboratory or Field Tracer Experiments. U.S. Salinity Laboratory, Riverside, CA, USA.
  • 31
    Hutzler NJ, Crittenden JC, Gierke JS, Johnson AS. 1986. Transport of organic compounds with saturated groundwater flow: Experimental results. Water Resour Res 22: 285295.
  • 32
    Brusseau ML, Jessup RE, Rao PSC. 1991. Nonequilibrium sorption of organic chemicals: Elucidation of rate-limiting processes. Environ Sci Technol 25: 134142.
  • 33
    Brusseau ML, Rao PSC. 1991. Influence of sorbate structure on nonequilibrium sorption of organic compounds. Environ Sci Technol 25: 15011506.
  • 34
    Zhu D, Hyun S, Pignatello JJ, Lee LS. 2004. Evidence for π–π electron donor–acceptor interactions between π-donor aromatic compounds and π-acceptor sites in soil organic matter through pH effects on sorption. Environ Sci Technol 38: 43614368.
  • 35
    Sithole BB, Guy RD. 1987. Models for teracycline in aquatic environment: II. Interaction with humic substances. Water Air Soil Pollut 32: 315321.
  • 36
    Gu C, Karthikeyan KG. 2005. Interaction of tetracycline with aluminum and iron hydrous oxides. Environ Sci Technol 39: 26602667.
  • 37
    Tanis E, Hanna K, Emmanuel E. 2008. Experimental and modeling studies of sorption of tetracycline onto iron oxides-coated quartz. Colloids Surf A 327: 5763.
  • 38
    Wershaw RL. 1999. Molecular aggregation of humic substances. Soil Sci 164: 803811.
  • 39
    Pils JRV, Laird DA. 2007. Sorption of tetracycline and chlortetracycline on K- and Ca-saturated soil clays, humic substances, and clay–humic complexes. Environ Sci Technol 41: 19281933.
  • 40
    Gu C, Karthikeyan KG. 2008. Sorption of the antibiotic tetracycline to humic–mineral complexes. J Environ Qual 37: 704711.

Supporting Information

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Additional Supporting Information may be found in the online version of this article.

FilenameFormatSizeDescription
etc_1726_sm_SupplData.doc20KSupplemental Data

Please note: Wiley Blackwell is not responsible for the content or functionality of any supporting information supplied by the authors. Any queries (other than missing content) should be directed to the corresponding author for the article.