Sorption of trichloroethene onto stylolites

Journal of Contaminant Hydrology 40 Ž1999. 1–23 www.elsevier.comrlocaterjconhyd

Sorption of trichloroethene onto stylolites Vera W. Langer a

a,),1

, Kent S. Novakowski b, Allan D. Woodbury

a

Department of CiÕil and Geological Engineering, 344 Engineering Building, UniÕersity of Manitoba, Winnipeg, MB, Canada R3T 5V6 b National Water Research Institute, Burlington, ON, Canada Received 22 October 1998; accepted 8 April 1999

Abstract Batch and double reservoir diffusion cell experiments are used to investigate sorption of trichloroethene onto stylolites. Stylolites are common features in carbonate rock formations, and might contain high amounts of organic matter. Due to the hydrophobic character of TCE, its transport in fractured carbonate aquifers could be significantly affected due to these aforementioned features. No research has been carried out to evaluate the impact of stylolites on organic pollutant transport. The main objectives of this experimental research are to verify TCE sorption onto stylolites, and to derive sorption and diffusion parameters describing the soluterrock interaction. Test results show that stylolites from the Lockport Formation in Southern Ontario, Canada contain significant amounts of organic carbon. Discrepancies are noted between carbon analyzer data and estimates from batch experiments and these might be due to TCE sorption also onto a clay mineral phase in stylolites or due to selective sampling. Adsorption and desorption behavior of TCE is investigated in specially designed double reservoir diffusion cells made out of stainless steel and Teflon. Three semi-analytical solutions for one-dimensional, reactive tracer migration through a porous medium are derived and used to evaluate TCE time–concentration profiles. Experimental data can best be modeled using a kinetic Langmuir sorption formulation with a maximum sorption capacity of 1.3 to 4.6 mgrg and a kinetic sorption constant of 4 = 10y7 to 5 = 10y7 lrmg sy1. TCE desorption into the exit reservoir is found to be a very slow kinetic process. No retardation is observed during TCE migration through a clay and organic matter free dolostone sample. TCE seems not to interact with calcareous mineral phases and moves conservatively. Bromide diffusion curves yield geometry factors Žg . for dolostone ranging between 0.05 and 0.13. From this study it can be concluded that TCE sorption is of importance when modeling TCE migration in fractured, stylolitic limestone aquifers where diffusion into the rock matrix takes place. Temporal TCE storage in rock matrix stylolites and fracture wall stylolites has to be taken

)

Corresponding author. Fax: q41-31-6314843; E-mail: [email protected] Present address: Rock Water Interaction Group, University of Bern, Dept. of Geol. and Min. and Petr., Baltzerstr. 1, 3012 Bern, Switzerland. 1

0169-7722r99r$ - see front matter q 1999 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 9 - 7 7 2 2 Ž 9 9 . 0 0 0 4 2 - X

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into consideration when evaluating actions of remediation at organic spill sites. q 1999 Elsevier Science B.V. All rights reserved. Keywords: Trichloroethene; Sorption; Stylolite; Diffusion; Organic carbon; Analytical solution

1. Introduction Trichloroethene ŽTCE. is a non-polar organic compound belonging to the group of highly volatile chlorinated hydrocarbons ŽHVCH.. It is highly toxic and is even at low concentrations a serious risk to human health ŽBedient et al., 1994; U.S. Environmental Protection Agency, 1994.. HVCHs are often considered to be immiscible with water but nevertheless have significant water solubility Ž0.15 to 8000 mgrl.. The US EPA Drinking Water Standard for TCE is 5 ppb Ž0.005 mgrl. ŽSpitz and Moreno, 1996.. Note its water solubility is 1100 mgrl at 258C ŽVerschueren, 1983., which is much higher than the aforementioned water standard. Therefore, small amounts of organic fuel leaking into our drinking water supply can represent a serious health risk and pollute a large water volume. In many carbonate aquifers around the world, groundwater has become contaminated by volatile organic compounds, through industrial activity or waste disposal practices. In most carbonate formations, the transport of pollutants to local water supplies or surface waters occurs primarily through a network of bedding plane and vertical fractures. It has been well established in the literature that sorption and matrix diffusion results in significant retardation of pollutants in fractured rock formations ŽFreeze and Cherry, 1979; McKay et al., 1993a,b; Novakowski and Lapcevic, 1994.. The amount of mass that will be stored temporary in the rock matrix will largely depend on the soluterrock interactions. It has been shown that in water, dissolved non-polar organic compounds will adsorb to the surface of organic particles present in geological material ŽKarickhoff et al., 1979.. The organic content in carbonate aquifers might be concentrated in stylolite layers. Stylolites are common pressure dissolution features and are found in carbonate formations throughout the world. Organic particles and other residual phases can become concentrated in these thin Ž0.1 to 5 mm., serrated dark layers during their evolution. Because stylolites are often associated with the presents of fractures, TCE sorption onto the organic fraction of stylolites may provide for a significant retardation mechanism. No studies to date have recognized the importance of stylolites on the transport of volatile, chlorinated hydrocarbons Žsuch as TCE. in fractured carbonates. The impact of stylolitization on hydrogeological issues, like contaminant transport, is unknown but is potentially significant. Because sorption of hydrophobic contaminants onto organic matter is thought to be reversible, stylolites might act as contaminant sinks during adsorption and as contaminant sources during desorption. Before attempting to simulate and predict the fate of organic pollutants in carbonate aquifers it is important to find parameters and mathematical expressions describing the soluterrock interactions accurately. It is common practice to utilize batch and diffusion experiments for these purposes. Diffusion experiments have been employed to investigate the effective diffusion coefficient, retardation factor, and effective porosity for

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tracers ŽShackelford, 1991; Bickerton, 1993; Novakowski and van der Kamp, 1996., and batch experiments have been used to determine the distribution coefficient Ž K d ., the fracture wall distribution coefficient Ž K dU ., the specific surface area Žh ., and the fraction of organic carbon Ž foc . when assuming linear-Freundlich sorption ŽKarickhoff et al., 1979; Chiou et al., 1981; Giger et al., 1983; Hassett and Banwart, 1989.. In order to evaluate the effective diffusion coefficient for a tracer, its free-water molecular diffusion coefficient Ž D 0 . and the rock matrix tortuosity factor Žg . have to be known. For geological materials, g ranges commonly from 0.01 to 0.5 ŽFreeze and Cherry, 1979.. The main objectives of this experimental and analytical study are Ži. to verify TCE sorption onto stylolites, Žii. to estimate the amount of organic carbon present in stylolites, Žiii. to determine sorption and diffusion parameters describing the soluterrock interaction, and Živ. to derive semi-analytical solutions describing kinetic mass transport behavior in the rock matrix. Assuming linear-Freundlich sorption, batch experiments are employed to estimate the amount of organic carbon in stylolites, and to evaluate the fracture wall distribution coefficient for stylolite lined fractures. In addition, diffusion experiments are conducted in horizontal, double-reservoir diffusion cells, which allows the simultaneous observation of adsorption and desorption behavior of reactive compounds. In a horizontal diffusion cell, the tracer is forced to migrate orthogonally through the linear stylolite layers. Analytical solutions for three different sorption models are derived in Laplace space and inverted using a numerical inversion algorithm. The models differ in their finite or infinite sorption capacity and kinetic sorption behavior. Each model is tested to explain and fit the experimental data. The geometric factor g for the rock matrix is determined using time–concentration profiles of the conservative tracer bromide. All rock samples are obtained from the Lockport Formation, which is a fractured dolostone aquifer with numerous, parallel-bedding stylolites at Smithville ŽOntario..

2. Stylolites in the Lockport Formation The term ‘stylolite’ is generic and does not imply a specific composition. Stylolitization takes place under pressure dissolution due to loading or tectonically related stresses. Carbonates are mobilized under high pressure, whereas relatively insoluble residual phases like silica minerals, ore minerals and organic matter are concentrated on internal discontinuities. Stylolites are classified following the geometric appearance of the line of discontinuity and their orientation ŽPark and Schot, 1968.. In the Lockport Formation, stylolites have a suture-like geometry with a large range in amplitude Ž0.1 to 10 mm.. The relationship to bedding planes is dominantly horizontal, with occasional stylolites inclined or in interconnected network form. In the Lincoln quarry near Smithville, individual stylolites extend laterally to over 100 m. Drill core observations associate two types of fractures with stylolitization; horizontal unloading fractures and vertical cracks. Vertical cracks are found most predominantly in the upper 25 m of the dolostone aquifer at Smithville. Horizontal unloading fractures are perpendicular to the maximum paleo-stresses and can be grouped into Ž1. coring induced

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mechanical breaks and Ž2. open bedding plane fractures. Unloading induced fracturing occurs in association with stylolites having low amplitude, leaving the stylolite horizontally divided into two parts. Statistical evaluations from drill cores obtained at the Smithville site verify a correlation between highly fractured zones and zones with numerous stylolites ŽRadcliffe, 1995.. The highest stylolite density of 2.9 stylolitesrm is found in the upper portion of the Lockport Formation. The organic content of stylolites and of the ‘pure’ rock matrix from the Lockport Formation was determined with a carbon analyzer on crushed and powdered material. As a result the organic carbon fraction in stylolites ranges between 0.01 and 5.26% and in the rock matrix between 0.0 and 0.29% Žunpublished data; Hilverda and Langer, 1997; personal communication.. These are significant amounts.

3. TCE sorption studies 3.1. Sampling procedure To conduct the batch and diffusion experiments, stylolitic dolostone samples are taken from the Lockport Formation in the vicinity of Smithville. For batch experiments, 11 stylolitic and 11 control dolostone samples without large, visible stylolites are cut from drill core. The complete drill core is stored in water basins to ensure saturation of the rock matrix at all times. The stylolitic dolostone samples are cut adjacent to potential bedding plane fractures so that one side exposed a stylolitic, dark gray to black surface. The control samples are cut next to the stylolitic dolostone samples. This sampling procedure assures that control samples are from the same depth and similar in chemical composition. It is desirable that the control samples should contain no organic carbon, but some of the controls show also dark lines and spots indicating possible organic matter Žorganic matter is also finely distributed throughout the rock matrix and therefore not visible.. The volume of each sample is measured beforehand, whereas the dry solid mass and porosity are calculated from weight loss by heating the samples in the oven at 1058C after equilibrium concentrations are determined. For the diffusion experiments, six thin Ž0.8 to 1.0 cm., water-saturated dolostone rock samples are cut parallel to the bedding. One sample contains no stylolite or other visible organic matter and is very light colored, whereas the other samples contain one or more stylolites. The stylolites embrace a large variety of amplitude Ž0 to 5 mm. and thickness Ž0.5 to 3 mm. ŽFig. 1.. The total porosity of the rock samples ranges from 2 to 8%. Samples are taken from different depth covering the upper 19 m of the Lockport Formation. 3.2. Batch experiments The batch experiments are performed in special 50 ml glass vials capped with a Teflon w-coated silicon septum in order to minimize volatile TCE loss. The glass vials and caps are EPA-certified for experimental use with volatile, organic compounds. The rock samples are put into the wide mouth glass vials and covered with 25 ml de-ionized

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Fig. 1. Stylolitic dolostone samples from diffusion cell experiments.

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H 2 O. A 500 ppb initial TCE solution is obtained by spiking the 25 ml H 2 O with 10 ml TCE stock solution Ž1250 grl. in methanol. After spiking, the glass vials are turned over to prevent any gas leakage through the caps. TCE concentrations in the solution are monitored using the Head Space Method ŽGarbarini and Lion, 1985., where 50 ml gas is extracted with a gas tight syringe from the head space and analyzed using a Photovac w 10 s Plus gas chromatograph. To account for volatile loss, all TCE concentrations are calibrated using three blanks. The samples are monitored over 19 days to assure equilibrium is reached. Throughout the experimental period less than 5% of the initial TCE mass are lost in the blank containers. 3.3. Diffusion experiments The double reservoir diffusion cells are constructed of stainless steel and Teflon w to assure no sorption of dissolved organic tracers onto the apparatus itself. Two stainless steel tubes are separated by the rock sample and are cone shaped towards the ends to prevent air entrapment during filling with de-aired, de-ionized water ŽFig. 2.. This design leaves only a small opening at either end, which are closed off by a Swagelog w fitting with a Teflon w-coated silicon septa in the cap. Through the septa, both reservoirs can be sampled with a syringe without opening the ends. Stainless steel tubes and rock samples are hold in place by a Teflon w coat, which closes tight around both reservoirs. This was accomplished by heating the Teflon w liner, which causes it to shrink. Six double reservoir diffusion cells and one blank single reservoir cell are build. The blank is capped at one side with a stainless steel plate and used to account for volatile TCE loss. The source reservoir in all cells has a volume of 20 ml. The exit reservoir is build with a volume of 20 ml for two cells and 209 ml for 4 cells. Having a 10-fold larger exit reservoir than source reservoir accelerates the drop in tracer concentration in the source

Fig. 2. Design of double reservoir diffusion cell.

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reservoir, but on the other hand, significantly prolongs equilibrium times and causes analytical problems due to very low tracer concentrations in the exit reservoir. Bromide ŽBry. is chosen as a conservative tracer because it does not naturally occur in the dolostone aquifer and therefore the background concentration in the saturated rock sample can be assumed to be zero. Bromide is thought to be non-reactive with calcareous minerals like calcite and dolomite as well as with organic matter ŽSmart and Laidlaw, 1977.. The free-water diffusion coefficient Ž D 0 . for a diluted potassium bromide electrolyte at 258C is 20.1 = 10y1 0 m2rs ŽStokes, 1950.. Although D 0 is a function of concentration, the experimental method is utilized over a sufficiently small concentration range so that D 0 can be considered constant. TCE, in contrast, is a volatile organic compound with a water solubility of 1100 mgrl at 258C. It reacts with organic matter in the dolostone samples by adsorption, as shown in the batch experiments. The TCE free-water diffusion coefficient is thought to be lower than bromide due to its larger molecule size, but is found to be similar. The source reservoirs of all double diffusion cells are filled with a 309 ppm Bry solution. After the source reservoirs are closed off, 20 ml liquid are extracted with a syringe through the septa and all diffusion cells are than spiked with 2 = 10 ml TCE stock solution. The stock solution is composed of 1000 mgrl TCE in methanol. This procedure results in an initial TCE concentration of 1000 ppb in all source reservoirs. The exit reservoirs remains filled with de-ionized water. Over a period of 6 to 7 months, 150-ml liquid samples are taken out of the source reservoirs with a gas tight syringe and analyzed for Bry and TCE. The liquid samples are injected in special, gas tight, 300 ml glass vials with a Teflon w coated silicon septa in its cap. After equilibration of liquid and gas phase, 50 ml gas phase from the glass vial is sampled and analyzed for TCE with a gas chromatograph Žhead space analysis.. In the first three sample rounds a Photovac w is used, in later rounds a 5890A Hewlett Packard gas chromatograph with a sensitivity below 1 ppb is used. For each sample round three TCE standards are prepared and analyzed for calibration. Subsequently, all TCE concentrations are calibrated through the TCE concentration in the blank diffusion cell to account for volatile loss. Throughout the experimental period about 10% TCE are lost in the blank diffusion cell. After TCE analyses, 100 ml liquid from the glass vial are diluted with 4 ml de-ionized water for Bry analyses with an ion chromatograph. Standards of 3, 5, and 10 ppm are used for calibration.

4. Analytical models for diffusion cell To classify TCE sorption behavior onto stylolites into linear or nonlinear, reversible or irreversible, instantaneous or kinetic behavior, analytical solutions for three different one-dimensional sorption models are derived in Laplace space Žsee Appendix A.. Solutions are obtained using an inversion algorithm by De Hoog et al. Ž1982.. Computed time–concentration profiles obtained from the analytical solutions are used to evaluate contaminant concentrations measured in source and exit reservoirs of the diffusion cells.

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Optimal sorption parameters are determined by calibration of the experimental and computed data. The % error is defined as N

% es

)

Ý Ž Ccomp .y Cexp . 1

2

)100.

N

Ý Ž Cexp .

Ž 1.

2

1

In a horizontal diffusion cell, advective flow and longitudinal dispersivity are zero. Therefore, a concentration gradient is the only driving force, and is given by Fick’s law for a conservative tracer. Adding a sorption term the governing equation becomes: EC Et

sD

E2 C Ex2

y

r b ECU u

Et

0FxGL

Ž 2.

with D s D0g

Ž 3.

where D: effective diffusion coefficient, D 0 : free-water diffusion coefficient, g : geometric factor related to tortuosity, C: solute concentration in the pore fluid ŽMrL3 ., CU : solute sorbed onto rock mass ŽMrM., u : porosity ŽLrL., r b : rock bulk density ŽMrL3 ., and ECrEt: change in concentration with time ŽM ly3 Ty1 .. Note that the x direction extends from 0 to L, where L is the rock sample thickness. For many dilute organic contaminants, the sorption process onto organic matter in the porous rock sample is thought to be linear and reversible ŽKarickhoff et al., 1979.. Assuming instantaneous equilibrium, the linear Freundich sorption isotherm can be applied with CU s CK d , where K d is the organic contaminant distribution coefficient. Transport Eqs. Ž4. and Ž5. are derived ŽFetter, 1993., Žsee also Table 1; models I and II.: EC Et

sD

E2 C Ex

2

y

r b EŽ KdC . u

Et

Ž 4.

or D E2 C

EC s Et

R Ex2

s DA

E2 C E x2

Ž 5.

with Rs1q

rb u

Kd

Non-equilibrium or kinetic sorption describes the interaction of the organic solute with the rock matrix where adsorption and desorption processes have different rates. Irreversible and reversible, linear, first-order kinetic sorption ŽFetter, 1993. is modeled using ECU Et

s k 2 C y k 3 CU Žsorption .

Ždesorption .

Ž 6.

Model

Tracer

Transport equation 2

Sorption isotherm 2 ..

Parameters evaluated

Conservative Žmodel I.

Bromide

ŽEC .rŽEt . s DŽŽE C .rŽE x with Ds D 0 g

Linear, reversible equilibrium Freundlich sorption Žmodel II.

TCE

ŽEC .rŽEt . s DŽŽE 2 C .rŽEx 2 .. yŽ r b ru .ŽŽEŽ K d C .. rŽEt .. Ž4. or ŽEC .rŽEt . s Ž Dr R .ŽŽE 2 C .rŽE x 2 .. Ž5. with Rs1qŽ r b ru . K d and DA s Ž Dr R . ŽEC .rŽEt . s DŽŽE 2 C .rŽE x 2 .. yŽ r b ru .Ž k 2 C y k 3 CU . Ž7.

CU sCK d

ŽECU rEt . s k 2 C y k 3 CU k 3 ™0 irreversible

k 2 ,k 3 : kinetic rate constants

ŽEC .rŽEt . s DŽŽE 2 C .rŽE x 2 .. yŽ r b ru . k 4 C Ž Cmax yCU . Ž9.

ŽECU .rŽEt . s k 4 C Ž Cmax yCU .

k 4 : kinetic rate constant Cmax : maximum sorption capacity

Irreversible and reversible, first-order kinetic sorption Žmodel III. Nonlinear, irreversible, kinetic Langmuir sorption Žmodel IV.

TCE

TCE

Formulations after Fetter, 1993.

D: effective diffusion coefficient g : geometry factor DA : apparent diffusion coefficient R: retardation coefficient K d : distribution coefficient

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Table 1 Transport equations for conservative and reactive tracers

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where k 2 , k 3 are rate constants Žmodel III.. The governing transport equation is: EC

E2 C

rb

Ž k 2 C y k 3 CU . Ž 7. Et u Ex Contaminant sorption might also be limited by the capacity of the organic material to adsorb. This nonlinear behavior can be modeled with the Langmuir sorption isotherm Žmodel IV.. The kinetic version is given by ECU s k 4 C Ž Cmax y CU . y k 5 CU , Ž 8. Et where k 4 and k 5 are rate constants and Cmax is the maximum sorption capacity ŽFetter, 1993.. Assuming that the desorption rate is near zero Ž k 5 ™ 0., contaminant transport is formulated through Eq. Ž9.. sD

EC

2

y

E2 C

rb

k C Ž Cmax y CU . Ž 9. Et u 4 Ex To solve Eqs. Ž5., Ž7. and Ž9. for contaminant migration in a horizontal diffusion cell the Laplace transform is applied. After obtaining a solution in Laplace space a semi-analytical approach is used to solve for C in the source or exit reservoir ŽBickerton, 1993.. sD

2

y

5. Data analyses and results from batch experiments and diffusion experiments TCE loss due to gas leakage andror degradation is eliminated from the results by calibration with control samples containing the same initial TCE concentration but no rock material. Due to the differences in organic content and rock mass in the samples, TCE concentrations drop at different time rates in batch and diffusion experiments. In batch experiments, sample 65E-ST3-20 shows the highest sorption capacity for TCE ŽFig. 3.. The largest amount of sorption takes place in the first two days. The concentration decreases to 0.26 CrC 0 , whereas in the following 17 days the concentration decline slows down and reaches equilibrium at 0.11 CrC 0 . In all other samples

Fig. 3. TCE time–concentration profiles from batch experiments for stylolitic dolostone and dolostone rock matrix.

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most of the sorption is completed after about eight days, and 50% from the total amount of sorption takes place in the first five days. The result from a batch experiment on ground, pure stylolitic matter is used for comparison Žunpublished data; Hilverda, 1997; personal communication.. The pure, ground stylolite material sorbed 75% of the starting 500 ppb Ž0.25 CrC 0 . in less than a day Ž20 h.. The five cm3 Ž11 g. pure, grounded

Fig. 4. Experimental and computed TCE and bromide time–concentration profiles for source and exit reservoirs. Sample 65E-ST18-1 is a light dolostone sample with presumably no organic matter and clay. Sample 65E-ST11-3 and 65E-ST1-17 are stylolitic dolostone samples. Model I simulates conservative mass transport, and model II accounts for linear, reversible Freundlich sorption during mass transport.

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stylolite material interacted much faster with TCE than most stylolitic dolostone samples. This behavior is expected because the grounded material exposed a larger surface area to the dissolved TCE. In the rock sample batch experiments, an initial mass of 12.5 mg TCE is injected. The stylolite mass is calculated to range from 1.7 to 5.2 g, and the TCE mass sorbed lays between 2.8 and 11.3 mg. The pure, ground stylolite sample had a mass of 11 g. The initial TCE mass injected was 17.5 mg, and the TCE mass sorbed was 14.6 mg. Diffusion experiments with stylolitic dolostone samples results in TCE time–concentration data, from source reservoirs, that are below their respective bromide data. With sample 65E-ST11-3 the normalized CrC 0 concentration decreases to about 0.2 and with sample 65E-ST1-17 to about 0.4 after 180 days ŽFig. 4.. This values are consistent with the different exit reservoir sizes, which cause different concentration gradients. In both diffusion cell exit reservoirs, the TCE concentrations remain low with 0.05 CrC 0 after 180 days, whereas Bry concentrations increase to the expected amounts as concentrations in the source reservoirs decrease. TCE and Bry solute concentrations in the control diffusion cell 65E-ST18-1 are almost identical. This is true for source and exit reservoirs.

6. Interpretations 6.1. Batch experiments Assuming a linear, equilibrium sorption isotherm, the TCE distribution coefficient for each sample is calculated using the ratio Ceq rCs . The organic carbon partition coefficient K oc Žl 3rM. is computed from the empirical, log–linear regression equation established by Karickhoff et al. Ž1979., and an octanol–water partition coefficient for TCE of log K ow s 2.29 ŽGiger et al., 1983.. log K oc s y0.21 q log K ow Ž Karickhoff et al., 1979 . The above relationship was established from batch experiments with different chlorinated hydrocarbons. The fraction of organic carbon Ž f oc . is estimated using the linear relationship K d s K oc f oc , where K oc for TCE is calculated from the above equation. Two main assumption are made using this simplified approach. First, K oc is independent of the type of organic matter, and second, adsorption onto organic matter is the only soluterrock interaction. As a result, K d ranges from 0.6 to 7.8 cm3rg Žaverage 2.6. in the stylolitic dolostone samples and between 0.4 and 2.0 cm3rg Žaverage 0.9. in the dolostone control samples ŽTable 2.. The data indicate clearly that the stylolitic rock matrix contains organic matter of significant amount contributing to TCE sorption. The fraction of organic carbon for all 22 rock samples ranges between 0.0032 and 0.066. Estimating the thickness of the stylolite layer with 0.5 to 1.5 mm on the rock sample surface and assuming a density of 2.2 grcm3, the stylolite mass is calculated for each sample. The stylolite mass from the 11 samples ranges between 1.7 and 5.2 g. Taking

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Table 2 TCE distribution coefficients for pure stylolite, stylolitic dolostone and dolostone rock matrix Type

Sample mass Žg.

TCE distribution coefficient K d Žcm3 rg.

Organic carbon f oc Ž%.

Stylolite pure Stylolitic dolostone Stylolite layer a Dolostone matrix

11 10.55–37.41 1.7–5.2 14.64–34.14

17.61 0.63–7.83 4.09–54.70 0.38–2.01

15Ž1. b 0.61–6.51 Ž11. 3.4–45.5 Ž11. 0.32–1.67 Ž11.

a

Estimating the stylolite mass in each sample and assuming an organic carbon background value equal to the dolostine matrix value, K d and f oc are calculated for stylolite layers. b Number in parentheses indicate the number of samples.

the fraction of organic carbon from the dolostone control sample as background value, f oc for the stylolite mass part is calculated ŽTable 2.. The average f oc for the stylolite layers is 0.1875 Žwith an average K d of 22.54 cm3rg. compared to a f oc of 0.15 Žwith K d of 17.61 cm3rg. found for the pure, ground stylolite sample in trial test-1. A wide range in K d for the different stylolite layers is expected due to the large variation in their composition. The dolostone matrix without stylolites has an average amount of 0.8% Ž f oc s 0.008. organic carbon. As stated above, fraction of organic carbon in stylolites are estimated on the assumption that TCE adsorbs only onto organic carbon. This cannot be verified at this point. Stylolites might also contain a large fraction of clay minerals which might contribute towards TCE sorption. This would result in lower f oc estimates; however, it would not diminish the important finding of TCE sorption onto stylolites. The fracture wall distribution coefficient is calculated for each stylolite surface by knowing the concentration of TCE in solution and the mass of TCE adsorbed per unit area of fracture surface. K dU s

mass of solute adsorbed per unit area of fracture surface Ž MrL2 . concentration of solute in solution Ž MrL3 .

Therefore, knowing K dU and K d , the specific surface area h for TCE sorption is evaluated and ranges between 1.9 and 8.2 cm2rg. The average specific surface area for TCE sorption onto stylolite fracture wall layer is 4.872 cm2rg. 6.2. Estimation of retardation factors Assuming linear-reversible equilibrium sorption, retardation factors ŽR. are estimated to range between 100 and 1000 for stylolite layers and between 10 and 50 for the rock matrix based on the range of TCE distribution coefficients and log K ow s 2.29 ŽFig. 5.. From 546 core sample measurements the average total porosity of the rock matrix in the Lockport Formation is 6.81%, and the bulk density is 2.51 grcm3. Estimation of the stylolitic fracture wall retardation factor of 30 to 400 is based on the range of TCE specific surface area data and an assumed fracture half aperture width Ž b . of 0.02 cm. K dU Rf s 1 q Ž 10 . b

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Fig. 5. Assuming linear, reversible equilibrium sorption, retardation factors for dolostone rock matrix and stylolite layers are estimated on the basis of the TCE octanol–water partition coefficient.

6.3. Diffusion experiments Curve fitting by calibration between analytical- and experimental data is used to evaluate the time–concentration data in source and exit reservoirs for both tracers. Decreases in solute concentration with time are strictly due to diffusion and sorption. For the conservative tracer bromide the semi-analytical solution from Bickerton Ž1993. is used. Knowing the initial concentrations in the source reservoir, porous medium and exit reservoir, and estimating the porosity and the diffusion coefficient, concentration– time data for the solute in source- and exit reservoir are calculated. Porosity values for the rock samples are determined after diffusion experiments are terminated. Bromide time–concentration profiles are used to evaluate the rock sample’s tortuosity related geometry factor g Žmodel I; Table 1.. The % error between computed and experimental data ranges from 2% to 7%. Curve fitting results in effective diffusion coefficients for Bry ranging from 0.9 = 10y10 m2rs to 2.6 = 10y10 m2rs ŽTable 3.. These values yield dolostone geometry factors between 0.045 and 0.13. TCE time–concentration data are evaluated employing the derived sorption models, and assuming that the free-water diffusion coefficient and the geometry factor for Bry and TCE are identical. First, all time–concentration profiles from source and exit reservoirs are analyzed assuming reversible, linear, equilibrium sorption Žmodel II. for TCE ŽFig. 5.. Experimental Bry and TCE concentrations in source and exit reservoir of control sample 65E-ST18-1 are best modeled with an effective diffusion coefficient of 2.61 = 10y1 0 m2rs and R s 1. This confirms that the free-water diffusion coefficient for TCE is very similar to the one for Bry and no adsorption or reaction takes place in the organic carbon free dolostone sample. Evaluating the TCE diffusion data from stylolitic dolostone samples, an apparent diffusion coefficient is found, which ranges between 5.0 = 10y1 0 and 12.1 = 10y10 m2rs ŽTable 3.. Separating retardation factors from the apparent diffusion coefficient, assuming D TCE equals Dy Br, results in unacceptable

Porosity Ž%.

Effective diffusion coefficient DU s D 0 g Žm2 rs.

g wx

model I q II: source reserÕoirs 65E-ST27-4 Bry TCE 65E-ST18-2 Bry TCE 65E-ST30-2 Bry TCE 65E-ST11-3 Bry TCE 65E-ST1-17 Bry TCE 65E-ST18-1 Žcontrol. Bry TCE

5.1 5.1 3.9 3.9 2.0 2.0 2.0 2.0 7.0 7.0 8.0 8.0

0.9=10y10

0.045 0.045 0.1 0.1 0.1 0.1 0.11 0.11 0.08 0.08 0.13 0.13

model I q II: exit reserÕoirs 65E-R1-17 Bry TCE 65E-R11-3 Bry TCE 65E-R18-1 Žcontrol. Bry TCE

7.0 7.0 2.0 2.0 8.0 8.0

Sample

Tracer

y10

2.0=10

y10

2.0=10

y10

2.2=10

y10

1.6=10

y10

2.6=10

1.6=10y10 2.2=10y10 2.6=10y10

0.08 0.08 0.11 0.11 0.13 0.13

Apparent diffusion coefficient DA s DU r R Žm2 rs.

Retardation factor Rs D 0 g r DA

5.4=10y1 0

0.17

5.6=10y1 0

0.36

5.0=10y1 0

0.4

12.1=10y1 0

0.18

5.0=10y1 0

0.32

2.6=10y1 0

1

1.0=10y1 2

160

2.0=10y1 2

110

2.6=10y1 0

1

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

Table 3 Results from diffusion experiments using models I and II

15

16

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

values between 0.2 and 0.4. Therefore, model II is not physically meaningful and must be rejected. However, the low TCE time–concentration data from exit reservoirs can be evaluated with model II, and result in retardation factors of 110 to 160. It is apparent that concentration profiles from the source and exit reservoirs cannot be modeled together when assuming reversible, linear, equilibrium sorption for TCE. A kinetic, irreversible sorption model Žmodel III. is much more appropriate to describe the observed behavior. This model leads to attenuation of the solute in the

Fig. 6. Experimental and computed TCE and bromide time–concentration profiles for source reservoirs. Rock samples are stylolitic dolostone. Model III simulates mass transport with irreversible, first-order kinetic sorption, and model IV accounts for nonlinear, irreversible kinetic Langmuir sorption during mass transport.

Model IV: source reservoirs Sample

Tracer

Porosity Ž%.

Effective diffusion coefficient a DU s D 0 g Žm2 rs.

g wx

Maximum adsorption capacity Cmax Žmgrg.

Kinetic rate constant k 3 Žlrmg sy1 .

65E-ST27-4 65E-ST18-2 65E-ST30-2 65E-ST11-3 65E-ST1-17 65E-ST18-1 Žcontrol.

TCE TCE TCE TCE TCE TCE

5.1 3.9 2.0 2.0 7.0 8.0

0.9=10y10 2.0=10y1 0 2.0=10y1 0 2.2=10y1 0 1.6=10y1 0 2.6=10y1 0

0.045 0.1 0.1 0.11 0.08 0.13

3.18 2.72 3.2 4.6 1.3 0

5=10y7 5=10y7 4=10y7 5=10y7 4=10y7 0

a

D 0 TCE s 20.1)10y10 Žm2 rs..

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

Table 4 Results from diffusion experiments using model IV

17

18

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

stylolite layers. In the given time frame the TCE desorption rate is zero, meaning k 3 s 0. The forward, first order kinetic constant k 2 is found to lie between 3.15 and 9.46 cmrg yeary1 . The steep drop in TCE concentration during the first 30 days suggests that the adsorption process is nonlinear and therefore concentration-dependent. However, TCE source reservoir concentrations in the two smaller diffusion cells Žsamples 65E-ST1-17 and 65E-ST27-4. seem to have reached equilibrium, which cannot be explained with a linear, irreversible sorption model ŽFig. 6.. In fact, a nonlinear, kinetic Langmuir type Žmodel IV. adsorption behavior with a zero desorption rate provides an even better correlation between experimental and analytical data for source and exit reservoirs. The % error decreased by about 10% ranging from 16 to 26% using a linear, irreversible sorption model Žmodel III. to about 6 to 15% using a kinetic Langmuir sorption model Žmodel IV.. The decrease in TCE concentrations, in the source reservoirs, depends on the rock sample sorption capacity. Because the organic content in all samples is assumed to be different, it is not surprised to observe a range of sorption capacities. Sample 65E-ST11-3 with the highest, visible organic content has also the highest TCE sorption capacity with 4.6 mgrg and sample 65E-ST1-17 with the thinnest stylolite layer has the lowest capacity with 1.3 mgrg ŽTable 4.. A forward kinetic rate constant Ž k 4 . ranging between 4 = 10y7 and 5 = 10y7 lrmg sy1 and a backward kinetic rate constant Ž k 5 . near zero fits the observed concentration profiles in source and exit reservoirs best.

7. Discussion and conclusions Using the semi-analytical solution for conservative tracers and knowing that the free-water diffusion coefficient for bromide in diluted systems is 20.1 = 10y1 0 m2rs, the tortuosity related geometry factor g for dolostone samples ranges between 0.045 and 0.13. This is in good agreement with published data for geological materials ŽFreeze and Cherry, 1979.. Bromide, as a conservative tracer, is used for comparison with TCE. Time–concentration profiles from both tracers show that when migrating through a clay and organic-matter free dolostone sample Ž65E-ST18-1. TCE moves conservatively. It can be concluded that no interactions between TCE and calcareous mineral phases take place. Benker et al. Ž1997. showed in a TCE-bromide tracer field experiment, conducted in a sand aquifer with less than 1% silt and clay and virtually no organic matter present, that TCE moves conservatively. TCE, as a non-polar organic compound, does not sorb onto quartz mineral surfaces. Note that the organic carbon fraction determined by carbon analyzer in stylolites is between 0.01 and 5.26% and in the rock matrix between 0.0 and 0.29%. Estimates of the organic carbon content in stylolites from batch adsorption experiments yields values between 3.4% and 45.5%. The organic carbon in the rock matrix is estimated with 1.67%. This discrepancy might be explained by TCE adsorption not only onto organic matter but also onto clay minerals in stylolites and rock matrix in batch experiments or simply by selective sampling. The small number of samples for carbon analyses might not represent the large compositional range of stylolites. Batch experiments and diffusion studies in this research clearly demonstrate the capability of stylolites to adsorb significant amounts of dissolved TCE. In the experi-

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

19

mental time frame of six months no significant amounts of desorbed TCE are measurable. Only 1.7% TCE from the initial mass injected into the source reservoir is detected in the exit reservoir. This supports experimental results from Culver et al. Ž1997. where only 5% of the initially sorbed mass from batch experiments had been desorbed after 600 h. The rate of desorption might also depend on long term TCE exposure ŽGrathwohl and Reinhard, 1993; Culver et al., 1997.. With large advective flows Ž24000 times the void volume., 72% of the initial soil TCE was removed in desorption column experiments from soils with 0.13% organic carbon ŽPavlostathis and Jaglal, 1991.. It seems that advective flow has a large influence on the time dependent desorption behavior of TCE. The difficulty of finding an appropriate distribution coefficient for TCE due to time dependent changes in mass transfer has to be pointed out. TCE distribution coefficients for stylolites from our adsorption batch experiments are found to range from 4.1 to 54.7 cm3rg, which correlates with published data ŽPavlostathis and Jaglal, 1991; Zytner, 1992; Culver et al., 1997.. Unusually low Ž1 cm3rg. TCE distribution coefficients for fine to medium grained soil were reported from Picatinny Arsenal ŽNJ. with 1.04% organic material ŽSahoo and Smith, 1997.. Retardation factors for stylolite layers are estimated from TCE distribution coefficients and range between 100 and 1000. Comparison between bromide and TCE concentration profiles in diffusion cell exit reservoirs yield retardation factors of 110 to 160 assuming linear, reversible equilibrium sorption. Batch experiments on stylolitic fracture surfaces reveal TCE specific surface areas of 1.9 to 8.2 cm2rg, which seems unusually high when compared to specific surface areas of peat moss Ž0.4 m2rg. and GAC Ž1300 m2rg., a commercial product with 74.1% organic carbon. Zytner Ž1992. was able to model TCE adsorption and desorption onto different organic soils with an organic carbon content ranging from 1.0% to 49.4% using a nonlinear, reversible Freundlich isotherm. TCE time–concentration profiles from our diffusion study does not support that finding. An interesting conclusion from Zytner’s Ž1992. batch experiments was that not only the adsorption potential increases with increasing organic carbon content, but also the soils retention potential. Due to unknown organic carbon content in the stylolite samples used in the diffusion cells this potential could not be evaluated. The nonlinear, kinetic Langmuir sorption model Žmodel IV. provides significantly improved fits of experimental TCE data with computed concentrations when compared to the reversible, linear equilibrium sorption model Žmodel II. and the irreversible, kinetic sorption model Žmodel III.. There appears to be a maximum sorption capacity of organic contaminant onto stylolites. The rate of mass transfer decreases as more solute adsorbs onto organic surfaces and time–concentration profiles flatten out. In other words, the TCE distribution coefficient is not constant. Stylolitic rock samples, containing different amounts of organic material, show a maximum TCE sorption capacity of 1.3 to 4.6 mgrg. The time dependent adsorption behavior is also seen in the batch experiment profiles. Depending on the stylolite rock sample Žits organic carbon content., equilibration times ranges between 8 and 17 days. The importance of stylolites on TCE transport through dolostone aquifers is apparent. An average of 55 horizontal stylolite layers with an average thickness of 0.5 mm per stylolite in the upper Lockport Formation ŽEramosa Member. can adsorb 78.7 to 278.3 g

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

20

TCE per square meter surface area. A dissolved TCE plume that extends over 5000 m2 could lead to TCE attenuation in stylolites of 393.25 to 1391.5 kg Ž270 to 955 l pure TCE. assuming that all stylolites are reached by the TCE plume and the maximum possible amount of TCE is adsorbed. Any bedrock remediation plan has to take into account that TCE desorption from stylolites back into the pore volume of rock matrix and into fractures is a very slow kinetic process. Instead of trying to retrieve dissolved TCE by pumping, enhancing in situ biodegradation might be in many cases the better option. Incorporating the determined physical and hydrogeological parameters into a larger scale model and a stylolite sensitivity analyses are the next steps in the ongoing research project.

Acknowledgements Most of the experimental work was carried out at the Canadian Center for Inland Waters in Burlington and funded through the Bedrock Remediation Project Smithville. Special thanks go to Kelly Miller and Lavinia Zanini for technical advise and help with chemical analyses. Analytical work continued at the University of Manitoba, and thanks are extended to Prof. Racz and Mr. Sarna from the Department of Soil Science and to Mr. Stainton from the Freshwater Institute. The authors like to acknowledge suggestions and comments from Prof. Benker and an anonymous reviewer, which helped to improve the paper significantly.

Appendix A. Application of Laplace transform The Laplace transform of f Ž C . is defined as `

cŽ x , p. s

H0 exp Ž ypt . C Ž x ,t . d t

Ž 11 .

with p: Laplace variable; c: solution of C in Laplace space. Eqs. Ž5., Ž7. and Ž9. transform to pc Ž x , p . s

D d2 c

Ž 12 .

R d x2

pc Ž x , p . s D

d2 c d x2

y

rb u

k2 c q

rb u

k3

CU p

Ž 13 .

and pc Ž x , p . s D

d2 c dx

2

y

rb u

k 4 c Ž Cmax y CU . , respectively.

Ž 14 .

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

21

Using the following initial and boundary conditions the particular solution of the Laplace transform Eq. Ž12., Ž13. and Ž14. are found: initial conditions: C Ž x ,0 . s 0 Cs Ž 0 . s C0

0FxFL

in the porous medium in the source reservoir

Ce Ž 0 . s 0

in the exit reservoir

boundary conditions: Cs Ž t . s C Ž 0,t . Ce Ž t . s C Ž L,t . The solution in Laplace space of Eqs. Ž12. – Ž14. is formulated as:

'

C0 Ž p y je Rp . exp cŽ x , p. s

x

ž' ž ' / / Rp

yl

D

D

'

C0 Ž p q je Rp . exp y

ž ž 'Rp

ly

x

'D

//

Ž 15 .

D with

D s Ž p y js'Rp .Ž p y je'Rp . exp Ž yl'Rp .

'

'

'

y Ž p q js Rp .Ž p q je Rp . exp Ž l Rp . and cŽ x , p.

s

ž

C0 y

s CU Ž pqt .

/

'

'

'

exp Ž y Ž pqt . l . Ž py je Ž pqt . . y Ž pq je Ž pqt .

.

ž

s CU Ž pqt .

/

D

=exp

ž(

Ž pqt .

ž

D

y C0 y q

x

/

s CU Ž pqt .

/

'

'

'

exp Ž Ž pqt . l . Ž pq je Ž pqt . . q Ž py je Ž pqt .

.

ž

s CU Ž pqt .

/

D

=exp

ž(

Ž pqt .

y

D

/ ž

x q

s CU p Ž pqt .

/

Ž 16 .

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

22

with A'D u

js s

Vs A'D u

je s

Ve L

ls

'D rb t s k2 u rb s s k3 u D s Ž p y js'p q t .Ž p y je'p q t . exp Ž yl'p q t .

'

'

'

y Ž p q js p q t .Ž p q je p q t . exp Ž l p q t . and cŽ x , p. C0 p y je p q t Ž Cmax y CU . exp

(

ž

/

s

p q t Ž Cmax y CU .

ž(

x

ž'

yl

D

//

D C0 p q je p q t Ž Cmax y CU . exp

(

ž

/

y

p q t Ž Cmax y CU . l y

ž(

ž

x

'D

//

D

Ž 17 . with

D s p y js p q t Ž Cmax y CU .

( = exp ž yl(p q t Ž C = ž p q j (p q t Ž C

ž

e

U

/ ž p y j (p q t Ž C y C . / y C . / y ž p q j (p q t Ž C y C . / y C . / exp ž l(p q t Ž C y C . / e

max

U

max

max

U

s

max

U

U

max

where, A: rock sample surface, Cs and Ce : concentration in the source and exit reservoir, respectively, and Vs and Ve : volume of the source and exit reservoir, respectively. The solution in real space was obtained using a numerical inversion algorithm ŽDe Hoog et al., 1982.. The sorbed concentration Ž CU . was updated after each time step. References Bedient, P.B., Rifai, H.S., Newell, C.J., 1994. Ground Water Contamination: Transport and Remediation. Prentice Hall PTR, New Jersey. Benker, E., Davis, G.B., Barry, A.D., 1997. Factors controlling the distribution and transport of trichloroethene in a sandy aquifer-hydrogeology and results of an in situ transport experiment. J. Hydrogeology 202 Ž1–4., 315–340.

V.W. Langer et al.r Journal of Contaminant Hydrology 40 (1999) 1–23

23

Bickerton, G.S., 1993. A Semi-Analytical Model for Solute Diffusion Trough Rock Samples of Finite Thickness. Unpublished BS Thesis, Univ. of Waterloo, Ontario, Canada. Chiou, C.T., Peters, L.J., Freed, V.H., 1981. Soil–water equilibria for nonionic organic compounds. Science 213 Ž8., 683–684, Washington. Culver, T.B., Hallisey, S.P., Sahoo, D., Deitsch, J.J., Smith, J.A., 1997. Modeling the desorption of organic contaminants from long-term contaminated soil using distributed mass transport rates. Environ. Sci. Technol. 31 Ž6., 1581–1587. De Hoog, F.R., Knight, J.H., Stokes, A.N., 1982. An improved method for numerical inversion of Laplace transforms, SIAM. Journal on Scientific and Statistical Computing 3 Ž3., 357–366. Fetter, C.W., 1993. Contaminant Hydrogeology, Macmillan, New York. Freeze, R.A., Cherry, J.A., 1979. Groundwater. Prentice-Hall, Englewood Cliffs, NJ. Garbarini, D.R., Lion, L.W., 1985. Evaluation of sorptive partitioning of nonionic pollutants in closed systems by headspace analysis. Environ. Sci. Technol. 19, 1122–1128. Giger, W., Schwarzenbach, R.P., Hoehn, E., Schellenberg, K., Schneider, J.K., Wasmer, H.R., Westall, J., Zobrist, J., 1983. Das verhalten organischer wasserinhaltsstoffe bei der grundwasserbildung und im grundwasser. Gas-Wasser-Abwasser 63 Ž9., 517–531, Zurich. ¨ Grathwohl, P., Reinhard, M., 1993. Desorption of trichloroethylene in aquifer material. Rate limitation at the grain scale. Environ. Sci. Technol. 36 Ž1., 89–108. Hassett, J.J., Banwart, W.L., 1989. The sorption of nonpolar organics by soils and sediments. Soil Science Society of America ŽSSSA. 22, 31–44, special publication. Karickhoff, S.W., Brown, D.S., Scott, T.A., 1979. Sorption of hydrophobic pollutants on natural sediments. Water Res. 13 Ž3., 241–248, Oxford Pergamon Press. McKay, L.D., Cherry, J.A., Gillham, R.W., 1993a. Field experiments in fractured clay till: 1. Hydraulic conductivity and fracture aperture. Water Resour. Res. 29 Ž4., 1149–1162. McKay, L.D., Gillham, R.W., Cherry, J.A., 1993b. Field experiments in fractured clay till: 2. Solute and colloid transport. Water Resour. Res. 29 Ž12., 3879–3890. Novakowski, K.S., Lapcevic, P.A., 1994. Field measurement of radial solute transport in fractured rock. Water Resour. Res. 30 Ž1., 37–44. Novakowski, K.S., van der Kamp, G., 1996. The radial diffusion method: 2. A semianalytical model for the determination of effective diffusion coefficients, porosity, and adsorption. Water Resour. Res. 32 Ž6., 1823–1830. Park, W., Schot, E.H., 1968. Stylolitization in carbonate rocks. In: Muller, G., Friedman, G.M. ŽEds.., Recent Developments in Carbonate Sedimentology in Central Europe. Springer, New York, NY, 66–74. Pavlostathis, S.G., Jaglal, K., 1991. Desorption behavior of trichloroethylene in contaminated soil. Environ. Sci. Technol. 25 Ž2., 274–279. Radcliffe, A.J., Analysis of Stylolitization and Fracturing in Rock Core from the Smithville Phase IV Site. Unpublished thesis, Univ. of Waterloo, 1995. Sahoo, D., Smith, J.A., 1997. Enhanced trichloroethene desorption from long-term contaminated soil using Triton X-100 and pH increases. Environ. Sci. Technol. 31 Ž7., 1910–1915. Shackelford, C.D., 1991. Laboratory diffusion testing for waste disposal — a review. J. Contam. Hydrol. 7, 177–217. Smart, P.L., Laidlaw, I.M.S., 1977. An evaluation of some fluorescent dyes for water tracing. Water Resour. Res. 13 Ž1., 15–33. Spitz, K.-H., Moreno, J., 1996. A Practical Guide to Groundwater and Solute Transport Modeling. Wiley-Interscience, New York. Stokes, R.H., 1950. The diffusion coefficients of eight uni-univalent electrolytes in aqueous solution. Am. Chem. Soc. J. 72, 2243–2247. U.S. Environmental Protection Agency. Technical background document to support rulemaking pursuant to the Clean Air Act — Section 122 Žg.. Ranking of pollutants with respect to hazard to human health. EPA-450r3-92-010. Emissions Standard Division, Office of Air Quality Planning and Standards, Research Triangle Park, NC, 1994. Verschueren, K., 1983. Handbook of Environmental Data on Organic Chemicals, 2nd edn. Van Nostrand Reinhold, New York, 1310 pp. Zytner, R.G., 1992. Adsorption–desorption of trichloroethylene in granular media. Water Air Soil Pollut. 65 Ž3–4., 245–255.