Environ. Sci. Technol. 2010, 44, 6242–6248
Assessment of Physical Leaching Processes of Some Elements in Soil upon Ingestion by Continuous Leaching and Modeling M I C H A E L P . B E E S T O N , * ,† A N D R E J P O H A R , ‡ JOHANNES T. VAN ELTEREN,† IGOR PLAZL,‡ Z D E N K A Sˇ L E J K O V E C , § M A R J A N V E B E R , ‡ AND HYLKE J. GLASS| National Institute of Chemistry (KI), Hajdrihova 19, SI-1001 Ljubljana, Slovenia, Faculty of Chemistry and Chemical Technology (FKKT), University of Ljubljana, Aˇskercˇeva 5, 1000 Ljubljana, Slovenia, Jozˇef Stefan Institute (JSI), Jamova 39, SI-1000 Ljubljana, Slovenia, and University of Exeter, Cornwall Campus, TR10 9EZ Penryn, Cornwall, U.K.
Received March 2, 2010. Revised manuscript received June 5, 2010. Accepted June 28, 2010.
The physical processes controlling the desorption of some elements (B, Cd, Co, Mn, Ni, and Sr) from soils in a continuous leaching system representing the human stomach are investigated here by fitting experimental leaching data to a mathematical particle diffusion model. Soil samples (50 mg) from Cornwall, UK, contained in a flowthrough extraction chamber (ca. 6.5 mL) were intimately contacted with artificial gastric solution at various flow rates (0.42-1.42 mL min-1) for up to ca. 4 h, followed by analysis of the fractions collectedwithinductivelycoupledplasmamassspectrometry(ICPMS). The leaching profiles of the various elements were fitted to a mathematical model incorporating two mass transfer processes (liquid film diffusion and apparent solid phase diffusion) to determine the effective external mass transfer coefficient(β)andtheapparentintraparticlesoildiffusioncoefficient (Da). A system of partial differential equations was solved numericallywithafinitedifferencediscretizationofthecomputational domain allowing the rate limiting physical desorption process(es) for each element to be determined. The (thermodynamic) driving force of the leaching process is definedbythedistributioncoefficient(Kd0)betweensoilandleachant. Although the Kd0 values investigated are very similar (ca. 6-15 L kg-1) for the elements studied with the exception of B (ca. 2.7 L kg-1), the leaching profiles are very different due to diffusionlimited processes. The elements may be classified as limited by β (B, Sr, and Cd), by Da (Co, and Mn) or by β and Da (Ni). This results in quantifiable parameters for the liability of elements in soil upon ingestion which may be implemented in future risk assessment protocols.
Introduction The dynamic measurement of the release of elements from soil has many well reported advantages such as reduction of * Corresponding author e-mail:
[email protected]. † National Institute of Chemistry (KI). ‡ University of Ljubljana. § Jozˇef Stefan Institute (JSI). | University of Exeter. 6242
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possible readsorption/redistribution (1), higher extraction speed (2), improved accuracy (3), more accurate representation of dynamic environments (4), lower sample and reagent requirements, lower risk of contamination and lower analyte loss, and the development of associations between elements (5–7). To realize the extent of these advantages, further understanding of the mechanisms (8, 9) and physical processes (3) controlling desorption of elements from soils during the measurement is required. This is particularly significant when the measurement is used for bioaccessibility studies of the gastric system, where material remains in the natural system for ca. 3-4 h (10). This limited time period generates nonequilibrium conditions hence preventing thermodynamic studies from drawing detailed conclusions regarding the mechanisms and physical processes of desorption. Diffusion processes control the release of elements from soil in a continuous leaching environment. These diffusion processes can be described by apparent solid phase diffusion (11), liquid film diffusion (3, 11), or both. The apparent solid phase diffusion contains effects due to other mechanisms like adsorption, soil-water partitioning and solubility along with soil composition and porosity (12). The apparent solid phase diffusion coefficient is affected by the pore size, the distribution of the pores, and the molar mass of the element in question. The liquid film diffusion quantifies the rate with which metals move from the surface of the particles, through the laminar film to the bulk liquid and is dependent on particle radius, bulk fluid composition, and effective molecular diffusion coefficient (13). The liquid film diffusion coefficient is also dependent on the hydrodynamic parameters such as flow rate. The kinetics of the release of elements from material within the human gastric system has been modeled previously. For example Ruby et al. (14), evaluated the mechanisms of Pb dissolution using first order, second order and parabolic diffusion models. This involved the use of a large (280 mL) reaction vessel with 20 g of soil that was sub sampled (3 mL) periodically. In an effort to approximate the human gastric system more accurately Gasser et al. (15), utilized continuous leaching to assess the kinetics of Pb released from mine waste. The extraction was completed under simulated gastric conditions using a stirred flow reactor and modeled empirically. In order to gain greater insight into the physical processes controlling element liability within the gastric system, the experimental procedure and modeling of the online extraction chamber previously described by Beeston et al. (5), was advanced. The model includes two mass transfer processes (liquid film diffusion and apparent solid phase diffusion) while equilibrium is governed by a distribution coefficient. Only a portion of the total concentration of each element in the soil is accessible to the extractant, as previously suggested by Van Elteren and Budicˇ (16). In addition, it is evident that, from the discrete sampling of the continuous leaching process, an elution profile can be developed. The focus of this work was to model the physical processes controlling desorption of elements from soils within the gastric system with the gastric phase of the physiologically based extraction test (PBET) (17). For this purpose a number of elements were selected (anionic and cationic) with different valency states. The effect of flow rate on the extraction was also investigated, as it is known, that an increase in the flow rate reduces the thickness of the liquid film about individual particles and hence increases the rate of extraction, therefore it is likely that an optimum flow rate exists. Given the 10.1021/es1006725
2010 American Chemical Society
Published on Web 07/16/2010
FIGURE 1. (A) Conceptualization of element leaching from a spherical soil particle within gastric solution under continuous leaching conditions. The concentration profiles within the particle are given at time (t) intervals for the extraction of the fractions (1/ 5th) of the available concentration (a0) of the element from the soil. (B) The concentration profile within the bulk solution representing the time intervals (arbitrary units) for the extraction of the fractions of a0 represented in Figure 1A (1/5th ) 5; 2/5th ) 15; 3/5th ) 60 and 4/5th ) 215). significance the gastric system, as a pathway through which contaminants can affect human health, samples for this investigation were collected from a populated region of Cornwall (UK), an area with historic mining activity. Model. The physical model outlined in Figure 1A is used to represent the mass transfer of elements from particles into solution. Analysis of the soil with X-ray fluorescence (XRF) defined the soil to contain, 67.3% SiO2, 2.54% K2O, 0.60% CaO, 0.90% TiO2, 0.17% MnO, 9.65% Fe2O3, 0.09% CuO, 0.09% ZnO, and 0.18% As2O5. The residual matrix is most probably of inorganic origin and composed of a mixture of oxides of the light elements that were not measured. Digestion of the sample with aqua regia and analysis with ICP-MS defined the soil to contain 3.98% Al2O3. Therefore based on this composition the analyzed radius of the particles is assumed to be constant during the extraction. Furthermore the elements are initially assumed to be homogenously distributed within a particle, an assumption that is regularly applied in models assessing diffusion in soils and aquifer materials (heterogeneous material) (18, 19). Diffusion in the particle can transport mass through the particle by solidstate diffusion through the bulk solid or liquid-phase diffusion within the pore structure of the particle (20). Within this investigation we assume both types of diffusion to be present; hence, the diffusion is referred to as “apparent intraparticle soil diffusion”. As a result of this assumption it is therefore assumed that the diffusion that occurs in the particles is
equal in all directions. Therefore the concentration distribution profile as a function of time of the elements within the particles is assumed to have an “onion” structure, with a uniform concentration at all points on each layer (radius). Diffusion within the stagnant laminar liquid film is assumed to be linear according to film theory. The movement of soil particles within all parts of the extraction chamber is characterized by intense chaotic movement, and hence the particles and solution are considered to be ideally mixed. As a result diffusion within the bulk solution is ignored. Finally the extraction is represented by the concentration profile in the bulk solution (Figure 1B). Note that the profile begins from an initial concentration of zero, with a sharp peak in the beginning. Model Development. As an ideally mixed system it is assumed that the concentration of the elements in question at the outlet of the extraction chamber is considered to be the same as the concentration inside any part of the extraction chamber at any given time. The model is based on the criteria that the particles are of identical size, which is based on analysis of the particles with SEM, and subsequent analysis of the SEM images with the image analysis software ImageJ (W.S. Rasband, U.S. National Institutes of Health, Bethesda, MD, 1997-2005). The methodology of modeling the system based on particles of an identical size is validated by previous work highlighting the fact that results obtained from a narrow and normal size distribution of particles is very close to the results obtained for particles of a uniform size (21). Also the assumption of particles of a spherical shape and uniform size has been assumed in a number of articles on the topic of sorption and desorption of elements from soils (e.g., refs 22–24). The procedural blank was subtracted from the sample data prior to the modeling, therefore the initial concentration of the gastric solution entering the extraction chamber could be defined as zero for the modeling purposes. The transient differential mass balance equation for the extraction chamber, which takes into account desorption of the elements into the mobile phase, can be written as follows: φv ∂c(t) 1 - ε ∂a j (t) c(t) + )0 ∂t εV ε ∂t
(
)
(1)
with the initial condition c(0) ) 0
(2)
where t is the time from the beginning of the experiment, φν is the volumetric flow rate, c is the concentration of the dissolved compound in the mobile phase (concentration per volume of mobile phase) and aj is the average concentration in the sorbent phase (concentration per volume of soil particles). ε is defined as the ratio of the void volume (V0) in the extraction chamber to its total volume (V), whereas 1 ε represents the fraction of solid particles in the system. The rate of desorption is also dependent on diffusion within each soil particle therefore the mass transfer inside the spherical particles was calculated simultaneously in order to obtain the concentration distribution within the particles. An implicit scheme was used to avoid numerical stability restrictions. For diffusion in a sphere the mass balance can be written as follows: Da ∂ 2 ∂a(r, t) ∂a(r, t) r ) 2 ∂t ∂r r ∂r
[
]
(3)
with the associated boundary conditions: a(r, 0) ) a0 VOL. 44, NO. 16, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
(4)
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∂a(0, t) )0 ∂r
(5)
Rβ ∂a(R, t) (c(t) - c*(t)) ) ∂r 3Da
(6)
where a is the element concentration in soil particles, r is the radius vector in particles, R is the average particle radius, c* is the equilibrium concentration of the element adjacent to the exterior surface of the particle and the interior (a) (see Figure 1A). Equation 6 represents the rate of desorption of the element and states that the diffusion flux for the external liquid film is equal to the intraparticle diffusion flux. Where β is the liquid film diffusion coefficient and Da is the apparent solid phase diffusion coefficient. The equilibrium concentration, c* was described by a linear (Henry type) isotherm and is therefore defined as: c* )
a1-ε Kd ε
(7)
where Kd [/] is the distribution coefficient of the element, defined by the ratio of the concentration of the element on the surface of the soil particle (per volume of mobile phase) to its equilibrium concentration adjacent to the particle in a volume of mobile phase. Assuming a linear adsorption isotherm, Kd was determined through a variable mass experiment (16) from the following expression:
Kdo )
a0 - c*
V m
c*
(8)
where a0 is total element concentration available to the gastric solution (per mass of soil). a0 is calculated according to Van Elteren and Budicˇ (16), whereas V/m is the ratio of the volume of the extraction chamber to the mass of soil, the distribution ratio Kd0 is reported in L kg-1. Kd [/] in eq 7 is obtained by the multiplication of Kd0 with the V/m ratio. Low Kd values result in a high extractability, whereas high Kd values result in a lower extractability under the same conditions. The mass transfer flux at the interface between the liquid and the particle is equal to the time rate of change of the average concentration of element inside the particle. The average concentration inside a spherical particle, needed for the evaluation of eq 1 is calculated from a known concentration distribution by the following integral: a j (t) )
3 R3
∫
R
0
a(r, t)r2dr
(9)
Sampling Method. The methodology used within this investigation involves the collection of fractions of solution over fixed time intervals which are independent of flow rate. The volumetric mass balance inside each vial can be written as follows: dV ) φν dt
(10)
and the elemental mass balance: d(cV) ) coutφν dt
(11)
where cout is the concentration of the element at the exit of the extraction chamber and the concentration entering the vial. The time dependent concentration of each element leaving the extraction chamber is calculated using eq 1, considering eqs 2-9. These concentrations represent the inlet concentrations, entering the vial for collection. With time, the vial fills to the final volume according to the sample time 6244
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and flow rate. The concentration inside the vial is therefore numerically calculated using eqs 10-11. The result represents the time-dependent concentration inside the vial (designated “vial concentration”). When the vial is replaced, the initial concentration in the vial is equal to the concentration at the extraction chamber outlet. The concentration inside the vial, prior to vial replacement, was therefore used to compare the numerical results with the experimental data and evaluate the leaching process.
Experimental Section Soil Characterization. Prior to the experimental procedure the soil was characterized through the use of X-ray fluorescence (XRF). Here ca. 0.5 g of material was prepared into a pellet using a hydraulic press. As primary excitation sources, the annular radioisotope excitation sources of Fe-55 (20 mCi, Isotope Products Laboratories, U.S.) and Cd-109 (10 mCi, Eckert & Ziegler, E.U.) were used. The emitted fluorescence radiation was measured by an energy dispersive X-ray spectrometer. For the analysis two detectors were used, the first being a Canberra Si(Li) detector equipped with a vacuum chamber, ADC (Canberra M8075) and spectroscopy amplifier (Canberra M2020) for the Fe-55 source. The second detector used was an Eg & G Ortec Si(Li) detector with an AFT research amplifier (Canberra M2025) and ADC (Canberra M8075) for the Cd-109 source. Spectrum from both detectors were collected by PC based MCA (S-100, Canberra). Additionally the soil was also characterized through analysis by ACME Laboratories, Vancouver, Canada. Here the digestion involved a 2:2:2 mixture of ACS grade concentrated HCl, concentrated HNO3, and demineralised water at >95 °C for 1 h in a water bath. The elemental concentrations were then determined with a Perkin-Elmer Elan 6000, ICPMS. Variable Mass Extraction. Varying amounts of each soil sample (0.05-0.25 g dry weight) were weighed into 15 mL “Sarstedt” tubes and 10 mL of the gastric solution prepared according to the PBET procedure (17) pH 2.5 was added to the tubes. The tubes were shaken horizontally in a thermostatted bath at 37 °C for 72 h (Julabo SW 22) using a shaking speed of 130 strokes per minute. The elemental concentrations within the supernatants were measured with ICP-MS (Agilent 7500ce) using the collision cell with helium gas to remove potential polyatomic interferences. The operating conditions were as follows: nebulizer, Babington; spray chamber, Scott, 4 °C; Rf power 1500 W; sampling depth, 6.9 mm; carrier gas flow rate, 0.9 L min-1; make up gas flow rate 0.15 L min-1; points per mass, 3; acquisition time, 14.74 s; and repetitions, 5. Extraction Chamber Preparation. A small amount (50 mg) of sample was accurately weighed into a standard 5 mL glass pipet with bulb, having a total volume (V) of ca. 6.5 mL, which is referred to as the extraction chamber in the remainder of the text. A Millipore Millex HV hydrophilic PVDF filter (0.45 µm pore diameter) caps the top of the extraction chamber. Cotton wool in the tube linked to the base of the extraction chamber prevents soil leaving the chamber, and Tygon tubing connects the base of the extraction chamber to a peristaltic pump (Gilson Minipuls Evolution). Prior to each experiment, all peristaltic (Tygon) tubes and connections were rinsed with 3% v/v HNO3, followed by Milli-Q, and dried. Sample Preparation. Contaminated soil taken from arable land located next to (
