Permeable Adsorbing Barriers for groundwater remediation from hexavalent chrome pollution
Michele Di Natale, Roberto Greco* and Dino Musmarra Dipartimento di Ingegneria Civile – CIRIAM, Centro di Ricerche in Ingegneria Ambientale, Seconda Università di Napoli, via Roma 29, 81031 Aversa CE, Italy Fax: +39 081 5037370 E-mail: [email protected] E-mail: [email protected] E-mail: [email protected] *Corresponding author
Abstract: Chrome ions removal from aqueous solutions by adsorption onto an activated carbon (Aquacarb 207EATM by Sutcliffe Carbon) and a South African coal char is illustrated. Different chrome concentrations, pH and salinity levels are tested. Experimental data are interpreted by means of a Langmuir approach-based model. Results show that a Permeable Adsorbing Barrier (PAB) made of Aquacarb 207EATM can protect a shallow aquifer from topsoil Cr(VI) leaching even during heavy rainfalls and that Cr(VI) is desorbed in dry periods, when clean water crosses the PAB, without any dangerous release in the aquifer.
Reference to this paper should be made as follows: Di Natale, M., Greco, R. and Musmarra, D. (2007) ‘Permeable Adsorbing Barriers for groundwater remediation from hexavalent chrome pollution’, Int. J. Environmental Technology and Management, Vol. 7, Nos. 1/2, pp.39–55.
Biographical notes: Michele Di Natale is Full Professor in Hydraulic and Maritime Works and Hydrology at the Engineering Faculty of Seconda Università di Napoli, where he is responsible for Hydraulic Structures and River Training courses. He is author of more than 100 scientific papers published in international and national journals, books and conference proceedings. His major areas of research are: waste water disposal and use of water resources for irrigation; unsteady groundwater flow; submerged jets and hydrodynamic actions on submerged pipelines; morphological evolution of mobile sea bed under the hydrodynamic actions of currents; diffusion of pollutants and their dispersion into natural water bodies; river mouth hydrodynamics; river training.
Roberto Greco received his PhD in Hydraulic Engineering in 1997. He is Associate Professor in Hydraulic and Maritime Works and Hydrology at the Seconda Università di Napoli, where he is responsible for Hydraulic Structures and Hydrology courses. He is author of more than 50 scientific papers published in international and national journals, books and conference proceedings and serves as reviewer for the Journal of Hydrology. His major areas of research are: infiltration in swelling and shrinking clay soils; development of innovative techniques for monitoring infiltration and
40 M. Di Natale, R. Greco and D. Musmarra
evaporation processes in unsaturated soils; modelling of interaction between waves and currents; modelling and monitoring of water supply networks; stochastic modelling of rainfall.
Dino Musmarra received his PhD in Chemical Engineering in 1989. He is Associate Professor in Chemical Plants at the Seconda Università di Napoli, where he is responsible of the Flue Gas Treatment and Waste Solid Treatment courses for the degree in Environmental Engineering. He is author of more than 100 scientific papers published in international and national journals, books and conference proceedings. His major areas of research are: flue gas desulphurisation, removal of mercury vapours from incinerator exhausted gases, adsorption of heavy metal ions from waste waters.
1 Introduction
Hexavalent chrome has been recognised as one of the most toxic inorganic water pollutants. Its high toxicity is mainly related to the capacity of its compounds to get accumulated in the aquatic food web, reaching human beings through the food chain and causing severe diseases in all life forms (Brauer and Wetterhahn, 1991). Hexavalent chrome compounds are released into the environment mainly from metal industry, agricultural activity and waste disposal. These emissions have dramatically increased over the last 20 years. The chrome compounds added to the environment eventually contaminate groundwater.
In order to reduce damages and harm to the environment and humans the 2000/60/CE directive considers hexavalent chrome as a primary hazardous pollutant and imposes the cessation or phasing out of discharges, emissions and losses before 2020. Meanwhile, two intermediate steps have been defined: the maximum value of total chrome in superficial water bodies has been fixed at 4 µg/l in fresh waters and 0.7 µg/l in marine waters in 2008 and 0.7 µg/l and 0.5 µg/l in 2015. Hence, new efforts in waste water remediation technologies are required to ensure compliance with the quality standards as well as pollutant capture from contaminated, groundwater bodies.
Various remediation technologies have been developed to clean up contaminated groundwater. Amongst these technologies, Permeable Reactive Barriers (PRBs) have been widely used over the last decade. A PRB is an in situ below-ground treatment zone of reactive permeable material that degrades or immobilises contaminants carried by groundwater flowing through it. Several full-scale applications have shown how PRBs may often represent a valid alternative to ex situ remediation techniques requiring groundwater pumping and aboveground treatment of contaminated water (Blowes et al., 2000). It is hardly possible to define general guidelines to choose between PRBs and Pump and Treat Systems (PTS), since many factors should be considered, such as the type of contaminant groups in the groundwater, the hydrogeology of the site, the rate of groundwater to be treated and the remediation goals. The results of a recent cost analysis, carried out by the US Environmental Protection Agency on 48 groundwater remediation sites in the USA and Europe (US EPA, 2001), show that PRBs’ capital costs are generally lower than PTS, especially when complex combinations of contaminants are present; that operating costs of PTS increase with the complexity of the treatment
Permeable Adsorbing Barriers for groundwater remediation 41
technology used; and that PTS capital and operating costs per unit volume of treated water are lower for systems treating more than 20 million gallons per year.
Two configurations are generally adopted for PRBs (Day et al., 1999):
• funnel-and-gate, with impermeable walls directing the contaminated plume towards the treatment zone
• continuous trench, intercepting the entire contaminated plume.
The use of reactive materials that are more conductive than surrounding soils ensures that groundwater spontaneously flows through the barrier. PRBs are best applied to shallow aquifers, since trenching machines do not allow digging barriers more than 10 m deep, while drilling and deep soil mixing techniques extend the reachable depth to 15 m. The use of other techniques, such as jet grouting, allowing for a depth of 30 m, is limited to small installations, due to high costs. More modern technologies, such as biopolymers slurry trenching, allow the installation of PRBs up to a depth of 25 m (Day et al., 1999).
Most of the installed PRBs are made of zero-valent iron (Fe0), either to remove heavy metals by redox reactions and precipitation (Benner et al., 1997, 1999; Blowes et al., 2000), or to degrade chlorinated hydrocarbons, sulphates and nitrates (Vogan et al., 1999; Cervini-Silva et al., 2002; Gandhi et al., 2002; D’Andrea et al., 2005). A weak point about Fe0 barriers is that accumulation of precipitates may limit barrier longevity by reducing porosity and conductivity (Blowes et al., 2000), with the creation of preferential flow paths through the barrier that reduce the contact time between the contaminated water and the reactive material (Kamolpornwijit et al., 2003). Current research is therefore investigating the potential use of alternative materials and physical chemical processes for PRBs (Benner et al., 2000, 2002; Guerin et al., 2002; Park et al., 2002; Ake et al., 2003; Kadirvelu et al., 2003; Komnitsas et al., 2004; Lee et al., 2004; Wan et al., 2004; Barton et al., 2004; Wantanaphong et al., 2005). A particularly promising physical process for groundwater remediation from heavy metals as well as chlorinated solvents is represented by adsorption (Stumm and Morgan, 1996). Adsorption phenomena include mass transfer from fluid bulk to a solid surface of adsorbing material and the formation of bonds between molecules in the fluid and atoms of the surface. Several materials exhibit good adsorption properties, such as zeolites (Czurda and Haus, 2002; Park et al., 2002; Woinarski et al., 2003), organic humic materials (Guerin et al., 2002; McGovern et al., 2002), byproducts of industrial processes (Smith et al., 2001; Komnitsas et al., 2004; Lee et al., 2004; Munro et al., 2004; Schipper et al., 2004; Wan et al., 2004), industrial or natural activated carbons (Lorbeer et al., 2002; Ake et al., 2003; Di Natale et al., 2005a, 2005b). The choice of an adsorbing medium is strictly related to the specific metal to be adsorbed: adsorption isotherms and adsorbed breakthrough curves must be experimentally determined (Lorbeer et al., 2002; Di Natale et al., 2005b).
The design of a PAB does not consist only in the choice of the absorbing medium; hydrodynamic and solute transport modelling of the contaminated aquifer is mandatory in order to know the pollutant plume spatial and temporal evolution (Gavaskar, 1999; Gupta and Fox, 1999; Eykholt et al., 1999). The geometrical design of a barrier is, therefore, a complex issue, since it depends on chemical process rate within the barrier, chemical and hydraulic characteristics, and site related factors such as pollution source extension and pollutant amount and concentration (Di Natale et al., 2005a).
42 M. Di Natale, R. Greco and D. Musmarra
The aim of this paper is to investigate the potential application of two adsorbing materials in PRBs for hexavalent chrome (Cr(VI)) removal from groundwater. The adsorption isotherms of the two materials are experimentally determined and a theoretical model is proposed for the interpretation of experimental results. A simple two dimensional flow example of application to a PRB is finally presented. Barrier performance is evaluated by numerical simulation of chrome transport and adsorption.
2 Experimental adsorption study
2.1 Materials
Two materials have been chosen in order to have an example of industrial sorbent (the Aquacarb 207EATM activated carbon) and a low cost precursor of activated carbons (the char of South African coal). The Aquacarb 207EATM is a commercially available non-impregnated granular activated carbon, provided by Sutcliffe Carbon. This carbon derives from a bituminous coal and its main chemical and physical characteristics have been reported in Di Natale (2004). This material has an average particle diameter of 1.2 mm. The char of South African coal used in the present work has been characterised by Arena et al. (1990). The average particle diameter is around 2.25 mm and its density is around 500 kg/m3.
2.2 Procedures
The experimental runs have been performed in a water jacketed thermo stated batch reactor by mixing a 200 ml aqueous solution, containing a given chrome concentration, with a known mass of sorbent. The chrome concentration in the solution ranged between 5 mg/l and 50 mg/l, while the sorbent mass ranged from 0.5 g to 10 g. The chrome solution was obtained by the dissolution of a hexavalent chrome salt, potassium dichromate (K2Cr2O7, reagent grade), in distilled water. The pH of the solution was adjusted by addition of nitric acid and potassium hydroxide, values of pH in the range 7–11 were tested. In this pH range the prevalent chrome ion is the hexavalent chrome.
All the experiments were carried out at ambient temperature (Φ = 25°C). Preliminary kinetic tests showed that a reaction time of 48 hours was required to reach equilibrium conditions.
To evaluate the equilibrium conditions, chrome concentration in the solution as well as on the carbon surface was measured. The equilibrium concentration of hexavalent chrome was measured by UV-VIS absorption spectrophotometry by means of complexation with diphenilcarbazide (APAT, IRSA-CNR, 2003). The amount of adsorbed chrome was measured by leaching the solid material with diluted nitric acid aqueous solution (HNO3, 1 M) which allows a complete desorption of chrome from the carbon surface. The solution was then analysed by means of Flame Atomic Absorption Spectrophotometry (AAS-F). A simple material balance made it possible to evaluate the amount of adsorbed chrome and to compare it with the measured one. The maximum admitted difference between measured and calculated adsorption capacity was set at ±8%.
Permeable Adsorbing Barriers for groundwater remediation 43
2.3 Adsorption isotherms
The experimental work has focused on the thermodynamic aspects of hexavalent chrome adsorption. In particular adsorption isotherms at pH 7 and 11 were realised to analyse the effect of pH. Experimental results (Figure 1) show that at pH = 7 the adsorption capacity of the Aquacarb 207EATM reaches the value of 6 mg/g in conjunction with the highest cCr concentration, while results obtained for the char are one order of magnitude lower at the same cCr concentration. At a higher pH (pH = 11) the adsorption capacity of both materials decreases to a few tens of micrograms per gram.
Figure 1 Adsorption isotherms of chromate ions as a function of pH. Comparison between experiments (symbols) and models (lines)
In order to give further details on the effect of pH on chrome capture, six solutions with the same volume (100 ml), the same concentration (25 mg/l) and the same mass of carbon (1 g), with a different initial pH were tested. Experimental results (reported in Figure 2 as the adsorption capacity as a function of the equilibrium pH) revealed that both the materials present a monotonic decrease of adsorption capacity at increasing pH.
44 M. Di Natale, R. Greco and D. Musmarra
Figure 2 Effect of pH on adsorption of chrome ions at initial concentration of 25 mg/l
Figure 3 presents the effect of solution salinity on the adsorption capacity of the Aquacarb 207EATM at pH 7 for a solution containing a chrome concentration of 20 mg/l. Experimental results show that the higher the salinity the lower is the adsorption capacity. The adsorption capacity becomes smaller by an order of magnitude by increasing the NaCl concentration up to 0.1 M.
Figure 3 Effect of solution salinity for a solution containing 20 mg/l of chrome at 25°C and pH = 8 for the Aquacarb 207EA activated carbon
Permeable Adsorbing Barriers for groundwater remediation 45
3 Adsorption theoretical model
Adsorption of ionic species in aqueous solution involves different elementary mechanisms by which carbon and ion can interact with each other. The first and most common mechanism considers the interaction between the ionic species in solution and the functional groups with acid/basic behaviour. The other involves the formation of bonds between electrophilic/nucleophilic sites and anions/cations, respectively. Electrophilic/nucleophilic sites are uniformly distributed over the surface of the adsorbing materials. The adsorption model, previously developed by Di Natale (2004) that studies the adsorption thermodynamics of chrome, mercury and arsenic ions, considers the presence of four different types of adsorption sites:
• surface functional groups with acid properties (σOH), which, by substitution reactions, may complexate the cations (σOM) (Benjamin et al., 2001)
• surface functional groups with basic properties (σOH), which may react directly with cations by addition reactions (σOHM) and, if protonated (σOH2
+), may also react with anions (σOH2A) (Benjamin and Leckie, 1981)
• electrophilic sites (σδ), which tend to react with anions and other nucleophilic substances in the solution (σδA–) (Alfarra et al., 2004)
• nucleophilic sites (σδ), with a more pronounced tendency to adsorb cations or other electrophilic substances (σδH+) (Alfarra et al., 2004).
The first two adsorbing sites are directly related to the presence of oxygen-carbon bonds on the carbon surface, mainly due to the activation process. The presence and the typology of these sites are well described by Boehm (2002). The last two typologies are related to the structure of the graphitic layer itself (Alfarra et al., 2004) and to the presence of other atoms on the carbon surface (e.g., sulphur, nitrogen, iron, etc.).
The cation M+ may be a metallic ion or the same H+; similarly, the anion, A–, may be the desired adsorbing anion or other inorganic or organic anions, such as OH– among the others. Thus, to describe the possible adsorption mechanisms involving cations or anions, one needs to consider also the possible parallel competitive adsorption pseudo-reactions which can occur with other similar ions present in solution, first of all H+ and OH–.
The first goal of this adsorption scheme is that it may simply explain the effect of pH on the adsorption of cations and anions reported in a large number of experimental works (e.g., Stumm and Morgan, 1996; Benjamin, 2002). Actually it is well known that the adsorption of anions is favoured by small pH levels while that of cations shows the opposite trend. According to this model the adsorption of cations (M+) on the carbon surface is connected with adsorption reactions which consider also the parallel reactions of the active sites with H+ or other cationic species in solution. Thus, at the same concentration of the desired cation (M+) a higher pH means less competition and thus a greater adsorption capacity of the sorbent. Similarly for the case of the anions in which the role of H+ is taken by OH–. Moreover, in this case it has to be kept in mind that the mechanism of adsorption on protonated basic sites may be significant only if a sufficient number of H+ ions are present in solution. Thus, it is strongly affected by the pH variation, as a higher pH means at the same time a higher competition with OH– and less protonated basic sites.
46 M. Di Natale, R. Greco and D. Musmarra
Moreover, the model may also provide a possible explanation of the effect of salinity, as a higher concentration of Na+ or Cl– ions means a stronger competition between these ions and the desired species.
Considering this reaction scheme it is possible to develop a descriptive model of hexavalent chrome adsorption which is based on the following hypotheses: validity of the Langmuir models; intrinsic exothermicity of adsorption phenomena (thus ω decreases by increasing temperature); adsorption of ionic species only; the adsorption of an ionic species does not depend on the presence of other ions; the acid/base activated sites are properties of the specific sorbent and do not depend on the adsorbate properties.
Thus, the proposed model requires a knowledge of the concentration of the ionic species in solution in equilibrium conditions. To evaluate these quantities, it is necessary to study the solution equilibrium under the experimental conditions at which the points of the adsorption isotherms have been measured. In this case, this is obtained by considering the mass balance on hexavalent chrome and on hydrogen, the conservation of electric charge and the relations representative of the equilibrium reactions (Table 1). In this study, the Davies’s procedure (e.g., Stumm and Morgan, 1986) for the evaluation of the activity coefficients of the ionic species has been used.
Table 1 Chromic acid hydrolysis reactions
Equilibria Log K
( )4 2 4HCrO H H CrO aq− ++ ⇔ 0.382
4 4CrO H HCrO− + −+ ⇔ –8.876 2
4 7 22HCrO CrO H O− −⇔ + 3.522
2OH H H O− ++ ⇔ 14.0
Source: Park and Jang (2002)
The analysis of equilibrium concentration in all the investigated conditions shows that the main ionic species in solution is by far CrO4
–2; therefore, the proposed adsorption model refers to CrO4
–2 only. The adsorption equation is:
0
* *4max max
0 *4
( )1 ( ) 1
H Cr CrH
H Cr OH OH ClH OH OH Cl
K c K c K cK c K c K c K c K c
δ
δδ δ δ
ω ω ω
ω ω+
+ − − −
= +
⋅= +
+ ⋅ + + +
(1)
where ω0 is the adsorption capacity related to the acid/base sites, ωδ is the one related to σδ sites, and Ki is the thermodynamic constant of the ith adsorption pseudo-reaction. In this equation, the only term which accounts for the presence of a given solution salinity is the competition with Cl–.
A regression analysis of the experimental data has been made with the software SigmaPlot® for non-linear regression analysis. The results are reported in Table 2 for both the materials. For the analysis of the experimental data on the activated carbon, the regression analysis also involved the acid/base sites, whose reaction parameters has been evaluated by Di Natale (2004). The absence of such sites on the carbon, due to the absence of an activation process, means that in this case only the adsorption reaction
Permeable Adsorbing Barriers for groundwater remediation 47
involving σδ sites has to be considered. In Figures 1–3 the prediction capacity of the model with respect to the experimental data is reported. As it can be observed, the model gives a good prediction of both the effect of pH (Figures 1 and 2) and solution salinity (Figure 3).
The mass transfer adsorption coefficient (kc) has been evaluated by using one of the numerous relationships available in the literature (Lancia et al., 1994):
0.45Sh 2 10.72Re p= + (2)
where Rep and Sh are the modified particle Reynolds and the Sherwood numbers, defined as:
Re1
Sh .
pp
c p
udn
k dD
ρµ
=( − )
=
(3)
Moreover, in equation (3) ρ is liquid density; u is liquid velocity; µ liquid viscosity; dp is average particle diameter; D is CrO4
–2 ion diffusivity; and n is bed porosity.
4 Example of application of an adsorbing barrier
As an example of the effectiveness of adsorbing barriers for polluted shallow aquifer remediation, the simple case of two dimensional groundwater flow and pollutant transport is numerically solved. In this hypothesis, dissolved chrome mass balance equation through a saturated porous medium may be written, with indicial notation, as follows (Bear, 1979):
48 M. Di Natale, R. Greco and D. Musmarra
2
0.Cr b Cr Crij
j
c u c cD St n t n x
ρ ω∂ ⋅∇ ∂∂+ + − + =∂ ∂ ∂
(4)
In equation (4) cCr represents chrome concentration in fluid [M/L3]; u is unit flux vector [L/T]; ω is chrome concentration on solid [M/M]; ρb is dry adsorbing material bulk density [M/L3]; n is soil porosity [L3/L3]; and S [M/(TL3)] represents a source of chrome concentration (i.e., leakage from topsoil). The components of the mechanical dispersion tensor Dij [L2/T] may be expressed as follows:
22
2 2
( ).
( )
y x yxL T L T
ijx y y x
L T L T
u u uuu u u
Du u u u
u u u
α α α α
α α α α
+ −
= − +
(5)
In equation (5), αL and αT represent, respectively, the longitudinal and transverse dispersivity coefficients [L].
Adsorption is supposed to take place by neglecting any interaction between chrome ions and other adsorbed species. Under isothermal conditions (Θ = 25°C), the second term on the left hand side of equation (4) reads:
[ * ( , )].bc Cr Cr ik a c c c
n tρ ω ω∂ = −
∂ (6)
In above equation a is particles external specific surface [L–1]; * ( , )Cr ic cω is adsorption isotherm [M/L3]; ci represents the concentrations of other dissolved ionic species, such as H+, OH–, Cl– [M/L3].
For the design of PABs, two requirements must be taken into account:
• capability to retain intense concentration peaks
• long-term performance.
For the first requirement to be addressed, contaminated flow travel time through the barrier should be long enough for adsorption process to take place. Therefore, barrier width W must satisfy the following inequality:
1( ) .cb
W k au
−> (7)
In the above equation ub represents groundwater flow velocity through the barrier. The second requirement may be addressed in two ways depending on whether the
contaminant source intermittent or continuous. For a continuous contaminant source, the barrier life Tb is defined by the total amount
of contaminants that, provided the concentration cb is of the incoming plume, the barrier is capable of adsorbing, and is given by:
bb
bbb Qc
WHcT
)(*ωρ= (8)
Permeable Adsorbing Barriers for groundwater remediation 49
In equation (8) ω*(cb) represents the concentration of solid in equilibrium with liquid phase concentration cb, H being the barrier height and Qb the flow rate through unit length of the barrier.
For intermittent contaminant sources, like all cases in which rainfall infiltration mobilises contaminants stored in the soil, the reversibility of the adsorption process implies that, when the PAB is crossed by clean groundwater, adsorbed species are removed from the solid matrix, giving rise to a contaminated plume at the exit of the barrier. A wide barrier ensures the release of adsorbed species to be gradual enough to avoid intolerable outgoing concentration, virtually extending barrier life indefinitely.
The case study of a contaminated soil plot, formerly occupied by an industrial plant, lying near a river bank, is now presented. High concentrations of Cr(VI) are present in the topsoil layer. Polluted leachate due to rainfall infiltration is carried into the neighbouring river by groundwater motion. A PAB is placed downstream within the aquifer to protect the river from pollution. The considered geometry is sketched in Figure 4.
Figure 4 Sketch of the geometry of the example of PAB application
For the study of such a test case, the initial liquid and solid chrome concentrations are assumed to be zero throughout the entire flow domain and the following conditions apply at flow domain boundaries:
inflow boundary 0 0 0 0
outflow boundary 0 0 0.
impermeable bed 0 0 0 0
soil surface 0 0 0
Cr
Cr Crx
Cr
Cr
c x y H tx
c cu x y H tt x
c x y ty
c x y H ty
λ
λ
λ
∂= = < < ∀ >
∂∂ ∂
+ = = < < ∀ >∂ ∂
∂= < < = ∀ >
∂∂
= < < = ∀ >∂
(9)
The numerical integration of equation (4) with conditions of equation (9) has been carried out by means of a first order finite differences implicit scheme.
The mean chrome concentration inside the contaminated topsoil layer of depth Hs = 1.0 m is assumed to be ωs = 10 µg/g. Assuming dry soil bulk density ρb = 1600 kg/m3, it follows that every square metre of the contaminated soil plot contains 16 g of Cr(VI) ions. Assuming a mean value of leachate concentration of 0.1 mg/l, it follows that in a moderately rainy climate it would take some centuries for the complete removal of Cr(VI) from the topsoil. Even in the assumed moderately rainy climate, rainfall events with some tens of millimetres of infiltration height in 24 hours may happen several
50 M. Di Natale, R. Greco and D. Musmarra
times every year. During such rainfall events, the concentration of the contaminated plume reaching the river is likely to assume dangerous values. As an example, Figure 5 shows the Cr(VI) plume concentration 1 week and 4 weeks after an extreme rainfall event, with an infiltration height of 100 mm in 24 hours and a concentration of Cr(VI) in the leachate c0 = 0.4 g/m3, as predicted by the numerical model. Groundwater concentration reaching the river largely exceeds the limit prescribed by the 2000/60/CE directive for surface waters.
Figure 5 Simulated Cr(VI) concentration [mg/l] distribution in the groundwater 1 week and 4 weeks after an extreme rainfall event with infiltration height of 100 mm in 24 h
The same rainfall event has been modelled in the presence of a PAB of 2.0 m width made of activated carbon Aquacarb 207EATM. Table 3 summarises all the characteristics of the considered example. Figure 6 shows the plots of groundwater Cr(VI) concentration with and without the barrier at various stages after the infiltration. The PAB ensures complete protection of the river from pollution, outgoing groundwater concentration being always below directive 2000/60/CE prescriptions.
The adsorbed chrome concentrations within the barrier are plotted in Figure 7. The maximum value attained during the whole phenomenon is less than 7.0 µg/g, which is far from the equilibrium concentration on solid, which, for the predicted liquid phase concentration flowing through the barrier is around 50 µg/g. Furthermore, only the upstream half of the barrier is initially affected by significant adsorption. As long as the incoming liquid concentration decreases, the adsorbed chrome distribution is progressively smoothened without releasing significant outgoing Cr(VI) concentration in the groundwater even three months after the event.
These results show how the PAB, also, is capable of protecting the river during a series of consecutive severe rainfall events, and that dry periods induce a slow clean up of the barrier with very low outgoing Cr(VI) concentration.
Permeable Adsorbing Barriers for groundwater remediation 51
Figure 6 Simulated Cr(VI) concentration [mg/l] in the groundwater at various stages after an extreme rainfall event with infiltration height of 100 mm in 24 hours in presence (PAB) and in absence of the PAB (no PAB)
Table 3 Summary of PAB application example parameters
Aquifer characteristics Aquifer bed depth, H 7.0 m Piezometric gradient, J 0.001 m/m Porosity, ns 0.4 m3/m3 Hydraulic conductivity, ks 0.001 m/s
Longitudinal dispersivity, αL 1.0 m
Transverse dispersivity, αT 0.2 m
Dry soil bulk density, ρs 1600 kg/m3
52 M. Di Natale, R. Greco and D. Musmarra
Table 3 Summary of PAB application example parameters (continued)
Topsoil characteristics Contaminated depth, Hs 1.0 m
Mean Cr(VI) concentration in soil, ωs 10 µg/g Event characteristics Infiltration height, h 100 mm Infiltration duration, Ts 24 h Leachate Cr(VI) concentration, cs 0.4 mg/l Leachate pH 7.0 PAB characteristics Adsorbing medium Aquacarb 207EATM activated carbon
Dry bulk density, ρb 600 kg/m3 Porosity, nb 0.4 m3/m3 Hydraulic conductivity, kb 0.001 m/s Barrier width, W 2.0 m Numerical model parameters
Horizontal space step, ∆x 0.2 m
Vertical space step, ∆y 0.2 m
Time step, ∆t 100 s
Figure 7 Simulated distribution of adsorbed Cr(VI) [mg/g] within the PAB at various stages after an extreme rainfall event with infiltration height of 100 mm in 24 h
Permeable Adsorbing Barriers for groundwater remediation 53
5 Conclusions
In this paper the use of a PAB as a tool for protecting groundwater from heavy metals pollution is investigated. An experimental study of the adsorption of chrome from aqueous solutions onto a two different adsorbing materials is presented. Such a study allows the identification of the best operating adsorbing conditions and the estimation of the Langmuir’s parameters. The results of the adsorption study were applied as a case study for the design of a PAB to protect a river from contaminated groundwater pollution.
The effectiveness of a PAB for the protection of a shallow aquifer is finally tested by means of numerical simulation of dissolved Cr(VI) transport and adsorption. The use of mathematical models for the design of a PAB is always necessary, since either the choice of adsorbing material or the definition of optimal barrier geometry depend on the hydraulic and chemical characteristics of the contaminated groundwater flow to be treated. Numerical results show how a PAB made of Aquacarb207EATM activated carbon is capable of capturing dissolved Cr(VI) ions leaching from contaminated topsoil during rainfall events. Although the reversibility of the adsorption process implies that Cr(VI) ions are released from the barrier when clean groundwater flows through it, simulations results show how, for a properly designed barrier, outflowing Cr(VI) concentration never attains dangerous levels.
Donnelly, K.C. and Phillips, T.D. (2003) ‘Porous organoclay composite for the sorption of polycyclic aromatic hydrocarbons and pentachlorophenol from groundwater’, Chemosphere, Vol. 51, pp.835–844.
Alfarra, A., Frackowiak, E. and Béguin, F. (2004) ‘The HASB concept as a mean to interpret the adsorption of metal ions onto activated carbons’, App. Surf. Sci., Vol. 228, pp.84–92.
APAT, IRSA-CNR (2003) in Belli, M., Centioli, B., de Zorzi, P., Sansone, U., Capri, S., Pagnotta, R. and Pettine, M. (Eds.): Metodi Analitici per le Acque, Vol. 1, pp.321–345.
Arena, U., Cammarota, A., Massimilla, L., Sicliano, L. and Basu, P. (1990) ‘Carbon attrition during the combustion of a char in a circulating fluidized bed’, Comb. Sci. Tech., Vol. 73, pp.383–389.
Barton, C.S., Stewart, D.I., Morris, K. and Bryant, D.E. (2004) ‘Performance of three resin-based materials for treating uranium-contaminated groundwater within a PRB’, J. Hazard. Mater. B, Vol. 116, pp.191–204.
Bear, J. (1979) Hydraulics of Groundwater, McGraw-Hill, New York, p.569. Benjamin, M.M. and Leckie, J.O. (1981) ‘Conceptual model for metal-ligand-surface interactions
during adsorption’, Environmental Science and Technology, Vol. 15, No. 9, pp.1050–1057. Benjamin, M.M. (2002) Water Chemistry, McGraw-Hill, New York. Benjamin, M.M., Kim, J., Kwan, P. and Chang, Y-J. (2001) ‘A novel ion exchange process for
arsenic removal’, AWWA 2001 Annual Conference, Washington DC. Benner, S.G., Blowes, D.W. and Ptacek, C.J. (1997) ‘A full-scale porous reactive wall for
prevention of acid mine drainage’, Ground Water Monit. Rem., Vol. XVII, No. 4, pp.99–107. Benner, S.G., Blowes, D.W., Ptacek, C.J. and Mayer, K.U. (2002) ‘Rates of sulphate reduction
and metal sulfide precipitation in a permeable reactive barrier’, Appl. Geochem., Vol. 17, pp.301–320.
54 M. Di Natale, R. Greco and D. Musmarra
Benner, S.G., Gould, W.D. and Blowes, D.W. (2000) ‘Microbial populations associated with the generation and treatment of acid mine drainage’, Chem. Geol., Vol. 169, pp.435–448.
Benner, S.G., Herbert Jr., R.B., Blowes, D.W., Ptacek, C.J. and Gould, D. (1999) ‘Geochemistry and microbiology of a permeable reactive barrier for acid mine drainage’, Environ. Sci. Technol., Vol. 33, pp.2793–2799.
Blowes, D.W., Ptacek, C.J., Benner, S.G., McRae, C.W.T., Bennett, T.A. and Puls, R.W. (2000) ‘Treatment of inorganic contaminants using permeable reactive barriers’, J. Contam. Hydrol., Vol. 45, pp.123–137.
Boehm, H.P. (2002) ‘Surface oxides on carbon and their analysis: a critical assessment’, Carbon, Vol. 40, pp.145–149.
Brauer, S.L. and Wetterhahn, K.E. (1991) ‘Chrome(VI) forms a thiolate complex with glutathione’, Journal of the American Chemical Society, Vol. 113, No. 8, pp.3001–3007.
Cervini-Silva, J., Larson, R.A., Wu, J. and Stucki, J.W. (2002) ‘Dechlorination of pentachloroethane by commercial Fe and ferruginous smectite’, Chemosphere, Vol. 47, pp.971–976.
Czurda, K.A. and Haus, R. (2002) ‘Reactive barriers with fly ash zeolites for in situ groundwater remediation’, Appl. Clay Sci., Vols. 1–2, pp.13–20.
D’Andrea, P., Lai, K.C.K., Kjeldsen, P. and Lo, I.M-C. (2005) ‘Effect of groundwater inorganics on the reductive dechlorination of TCE by zero-valent iron’, Water Air Soil Poll., Vol. 162, pp.401–420.
Day, S.R., O’Hannesin, S.F. and Marsden, L. (1999) ‘Geotechnical techniques for the construction of reactive barriers’, J. Hazard. Mater., Vol. 67, pp.285–297.
Di Natale, F. (2004) Adsorbimento di Ioni Metallici in Soluzione Acquosa; Applicazioni al Caso di Cromo, Arsenico e Mercurio, PhD Thesis, DIC, University of Naples, Napoli, Italy.
Di Natale, M., Greco, R., Molino, A. and Musmarra, D. (2005a) ‘Design of permeable adsorbing barrier to mitigate hydraulic risk by chrome contaminated groundwater’, Proceedings of XXX IAHR Congress, Seoul, pp.2322–2331.
Di Natale, M., Greco, R., Karatza, D. and Musmarra, D. (2005b) ‘Groundwater remediation by using permeable adsorbing barrier to remove heavy metals’, Proceedings of XXX IAHR Congress, Seoul, pp.5326–5334.
Eykholt, G.R., Elder, C.R. and Benson, C.H. (1999) ’Effects of aquifer heterogeneity and reaction mechanism uncertainty on a reactive barrier’, J. Hazard. Mater., Vol. 68, pp.73–96.
Gandhi, S., Oh, B-T., Schnoor, J.L. and Alvarez, P.J.J. (2002) ‘Degradation of TCE, Cr(VI), sulfate, and nitrate mixtures by granular iron in flow through columns under different microbial conditions’, Water Res., Vol. 36, pp.1973–1982.
Gavaskar, R. (1999). ‘Design and construction techniques for permeable reactive barriers’, J. Hazard. Mater., Vol. 68, pp.41–71.
Guerin, T.F., Horner, S., McGovern, T. and Davey, B. (2002) ‘An application of permeable reactive barrier technology to petroleum hydrocarbon contaminated groundwater’, Water Res., Vol. 36, pp.15–24.
Gupta, N. and Fox, T.C. (1999) ‘Hydrogeologic modeling for permeable reactive barriers’, J. Hazard. Mater., Vol. 68, pp.19–39.
Kadirvelu, K., Kavipriya, M., Karthika, C., Radhika, M., Vennilamani, N. and Pattabhi, S. (2003) ‘Utilization of various agricultural wastes for activated carbon preparation and application for the removal of dyes and metal ions from aqueous solutions’, Bioresource Technology, Vol. 87, No. 1, pp.129–132.
Kamolpornwijit, W., Liang, L., West, O.R., Moline, G.R. and Sullivan, B.A. (2003) ‘Preferential flow path development and its influence on long-term PRB performance: column study’, J. Contam. Hydrol., Vol. 66, pp.161–178.
Komnitsas, K., Bartzas, G. and Paspaliaris, I. (2004) ‘Efficiency of limestone and red mud barriers: laboratory column studies’, Miner. Eng., Vol. 17, pp.183–194.
Permeable Adsorbing Barriers for groundwater remediation 55
Lancia, A., Musmarra, D., Pepe, F. and Volpicelli, G. (1994) ‘Characteristic times for limestone particle distribution in the production of gypsum from the wet flue gas desulfurization process’, Env. Sci. Tech., Vol. 28, pp.1031–1036.
Lee, T., Park, J-w. and Lee, J-H. (2004) ‘Waste green sands as reactive media for the removal of zinc from water’, Chemosphere, Vol. 56, pp.571–581.
Lorbeer, H., Starke, S., Gozan, M., Tiehm, A. and Werner, P. (2002) ‘Bioremediation of chlorobenzene-contaminated groundwater on granular activated carbon barriers’, Water Air Soil Poll. Focus, Vol. 2, pp.183–193.
McGovern, T., Guerin, T.F., Horner, S. and Davey, B. (2002) ‘Design, construction and operation of a funnel and gate in situ permeable reactive barrier for remediation of petroleum hydrocarbons in groundwater’, Water Air Soil Poll., Vol. 136, pp.11–31.
Munro, L.D., Clark, M.W. and McConchie, D. (2004) ‘A bauxsol-based permeable reactive barrier for the treatment of acid rock drainage’, Mine Water Env., Vol. 23, pp.183–194.
Park, J-B., Lee, S-H., Lee, J-W. and Lee, C-Y. (2002) ‘Lab scale experiments for permeable reactive barriers against contaminated groundwater with ammonium and heavy metals using clinoptilolite (01-29B)’, J. Hazard. Mater., Vol. 95, pp.65–79.
Park, S-J and Jang, Y-S. (2002) ‘Pore structure and surface properties of chemically modified activated carbons for adsorption mechanism and rate of Cr(VI)’, Journal of Colloid and Interface Science, Vol. 249, No. 2, pp.458–463.
Schipper, L.A., Barkle, G.F., Hadfield, J.C., Vojvodic-Vukovica, M. and Burgess, C.P. (2004) ‘Hydraulic constraints on the performance of a groundwater denitrification wall for nitrate removal from shallow groundwater’, J. Contam. Hydrol., Vol. 69, pp.263–279.
Smith, C.C., Anderson, W.F. and Freewood, R.J. (2001) ‘Evaluation of shredded tyre chips as sorption media for passive treatment walls’, Eng. Geol., Vols. 1–4, pp.253–261.
Stumm, W. and Morgan, J.J. (1996) Aquatic Chemistry, 3rd ed., Wiley & Sons, New York. US EPA (2001) Cost Analyses for Selected Groundwater Cleanup Projects: Pump and Treat
Systems and Permeable Reactive Barriers, Revised Version, EPA 542-R-00-013. Vogan, J.L., Focht, R.M., Clark, D.K. and Graham, S.L. (1999) ‘Performance evaluation of a
permeable reactive barrier for remediation of dissolved chlorinated solvents in groundwater’, J. Hazard. Mater., Vol. 68, pp.97–108.
Wan, M-W., Petrisor, I.G., Lai, H-T., Kim, D. and Yen T.F. (2004) ‘Copper adsorption through chitosan immobilized on sand to demonstrate the feasibility for in situ soil decontamination’, Carbohydr. Polym., Vol. 55, pp.249–254.
Wantanaphong, J., Mooney, S.J. and Bailey, E.H. (2005) ‘Natural and waste materials as metal sorbents in permeable reactive barriers (PRBs)’, Environ. Chem. Lett., Vol. 3, pp.19–23.
Woinarski, A.Z., Snape, I., Stevens, G.W. and Stark, S.C. (2003) ‘The effects of cold temperature on copper ion exchange by natural zeolite for use in a permeable reactive barrier in Antarctica’, Cold Regions Sci. Technol., Vol. 37, pp.159–168.
