transport of phthalate-esters in undisturbed and unsaturated soil columns

Journal of Contaminant Hydrology, 8 (1991) 111-133 111 Elsevier Science Publishers B.V., Amsterdam

Transport of phthalate-esters in undisturbed and unsaturated soil columns

T. Zurm/ihl, W. Durner and R. Her rmann

Department of Hydrology, University of Bayreuth, P.B. 101251, 8580 Bayreuth, Germany

(Received August 29, 1990; revised and accepted April 16, 1991)

ABSTRACT

Zurmiihl, T., Durner, W. and Hermann, R., 1991. Transport of phthalate-esters in undisturbed and unsaturated soil columns. J. Contam. Hydrol., 8:111-133.

Breakthrough curves (BTC's) of five phthalate esters (PE's) and chloride as a tracer were measured during unsaturated, steady water flow conditions on two undisturbed soil columns with different organic carbon contents (OC). Two conceptual models, the local equilibrium assumption (LEA) model and the physical nonequilibrium (PNE) model, were used to simulate the observed BTC's. The unknown parameters for the two models were estimated either by fitting model predictions to the experimental data, or, if possible, by independent measurements. Results of the tracer BTC's indicate that for the soil with the lower organic content the local equilibrium assumption was valid, whereas for the more aggregated soil, with higher OC, physical nonequilibrium existed. For the PE's an incomplete mass-balance was observed which increased with increasing hydrophobicity of the PE's and which was greater for the soil with the higher OC. Since microbiological degradation of the PE's was hindered, it cannot be the major cause for this mass deficit. Consequently, it is partly attributed to a chemical nonequilibrium, which is due to slow diffusion of the contaminants into the organic matrix of the soil and a possible subsequent entrapment of these substances in the organic matter. When modeling the PE BTC's, it was necessary to introduce a_ non-zero degradation term to account for this nonequilibrium. For the soil with the lower OC the LEA-model described well the measured BTC's for the PE's. When independently measured parameter values for the other soil were used, the LEA-model failed to describe any of the observed BTC's whereas good agreement between measured and calculated data was reached with the PNE-model.

INTRODUCTION

The protection of groundwater as the major drinking water resource in Germany requires the assessment of potential pathways of contamination. Applications of sewage sludge on soils is recognized as a potential source of groundwater contamination. The amount of sewage sludge will increase steadily in the future as more sewage treatment plants will be equipped with phosphate elimination steps. The use of sewage sludge as a soil fertilizer is

0169-7722/91/$03.50 © 1991 Elsevier Science Publishers B.V. All rights reserved.

112 1-. ZURMt~nL ET AL.

more preferable than its combustion or its deposition in landfills. It must be certain, however, that hazardous substances neither accumulate in the soil or in plants nor pose a threat to the groundwater.

Whereas transport and accumulation of heavy metals in soil after applications of sewage sludge have been subject to a series of studies (e.g., Lake et al., 1984; Davis et al., 1987; Fletcher and Beckett, 1987), there is far less known on the behaviour of organic pollutants. A relevant group of organic contaminants in sewage sludge are phthalate esters (PE's) (Zurm/ihl, 1990). PE's readily undergo microbiological degradation (Kurane, 1986), but because of the large amounts released and because of their toxicological relevance (Giam et al., 1984) PE's are included in the EPA list of priority pollutants (Callahan et al., 1979).

In order to quantitatively assess the potential of organic substances to contaminate groundwater, it is necessary to recognize the relevant processes, to describe the behaviour in a valid model, and to determine the necessary model parameters. Parameter determination directly in the field is certainly preferable but is often either not possible or not practical. Hence, parameter determination requires laboratory experiments.

A common means of investigating the transport behaviour of chemicals is the measuring of breakthrough curves (BTC's) of contaminants during miscible displacement in soil columns. By modeling the BTC's in combination with the use of weighted least squares optimization procedures (Kool et al., 1987) it is possible to determine parameters which are not accessible to direct measurement. Most models that have been used to describe solute transport by miscible displacement have been based on the convective-dispersive transport model, which assumes local physical and chemical equilibrium. Experimental observations, however, have often differed from the sigmoidal BTC's predicted by these models. Physical nonequilibrium processes, i.e., water flow in different domains (Van Genuchten and Wierenga, 1976), as well as chemical nonequilibrium processes, i.e., sorption kinetics (Selim et al., 1976; Cameron and Klute, 1977), are thought to be responsible for asymmetric BTC's. Nkedi-Kizza et al. (1984) demonstrated the mathematical equivalence of a physical and chemical nonequilibrium model. Therefore, parameter fitting alone cannot be used to differentiate chemical or physical effects on transport behaviour. Furthermore, most of the studies on transport of organic chemicals in soils are carried out in disturbed (Schwarzenbach and Westall, 1981), and/or water saturated (Rao et al., 1979; Nkedi-Kizza et al., 1983) soil columns.

Any transfer of results from laboratory studies to the field is subject to a series of problems. For example, disturbed samples neglect the influence of soil structure on pore geometry. Further, it appears preferable to make measurements under flow conditions as they may occur in nature.

TRANSPORT OF PHTHALATE-ESTERS 113

Therefore, the aim of this s tudy was to identify the processes and to partially validate models for the t ransport of five phthalate esters in naturally- structured soil cores under unsatura ted water flow conditions. To date no investigations have been reported on the distr ibution and t ransport behaviour of phthala te esters in undis turbed soil samples.

T H E O R Y

Transport equations

The convective-dispersive t ransport of a sorbing solute during one-dimen- sional steady water flow through a porous med ium subject to first-order decay both in dissolved and adsorbed phase is given by:

p Os Oc 02c Oc p 00--t + 0--}- = D ~ - V~z z -- #,c - / t s 0 S (1)

where c is the concentra t ion in solution (M L-3), s is the adsorbed concen- trat ion per unit mass of solid phase (M M-~), D is the dispersion coefficient (L: T-~ ), v is the average pore-water velocity (L T - 1 ), p is the soil bulk density (M L-3), 0 is the volumetric water content (L 3 L-3),/~t a n d / t s are the first- order decay rate constants for the liquid and solid phases of the soil, respectively (T-~), t is t ime (T) and z is the distance (L).

Assuming adsorpt ion to be instantaneous and to be described by a linear isotherm, a dimensionless retardat ion coefficient, R, can be defined as:

R = 1 + p K p O - 1 (2)

where Kp (M L -3) is the distr ibution coefficient (Kp = s . c- l ) . Substi tut ing (2) in (1) yields:

~c 02c Oc R 0-t = P t?z ---5 - v ~z - #c (3a)

where the new rate coefficient of decay, # (T- l ) , is given by:

It =- IAl + f l spK pO -1 (4a)

I f the two degradat ion coefficients are identical (Pl = #s = #*) eq. 4a reduces t o :

p = Rp*

In dimensionless form eq. 3a can be written as

0c 1 (~2C 0C R ~?T -- P O x 2 Ox ROc

where T = vt/L, P = vL/D, x = z /L and 0 = #*L/v.

(4b)

(3b)

114 r. ZURMOHL ET AL.

For a semi-infinite system and a pulse injection of a chemical, eq. 3b can be solved analytically, subject to the following initial and boundary conditions (flux-averaged concentrations, Parker and Van Genuchten, 1984a):

c(x, 0) = 0 (Sa)

~c (~ , T) 0 (5b) dx

{;0 f°r 0 < T~< T0 c(0, T) = (5c)

for T > To

where c o is the input concentration, which is applied during the pulse period 7"0.

The analytical solution of eq. 3 subject to eq. 5 is given by Parker and Van Genuchten (1984b). Equation 3 will hereafter be referred to as the local equilibrium assumption (LEA) model.

If the porous medium is assumed to contain a mobile water phase, in which the advective-dispersive transport of solutes occurs, and an immobile water phase, with which the solutes can exchange, then the following equations result (Van Genuchten and Wagenet, 1989):

(~Cm = OmOm (~2cm 8Cm (0 m + fpKp) Ot Oz -'--T - vmOm Oz

- - ~(Cm - - t im) - - Cm(0ml/Im + fPgpl~sm) (6)

OCi m [0im + (1 - - f ) p g p ] ~t = 0~(cm - - cim) - - Cim[0imfllim "~- (1 - - f ) p g p I 2 s i m ]

(7)

where the subscripts m and im refer to the mobile and immobile water phase, respectively, f represents the fraction of the sorption-sites for which equilibrium is instantaneous with the mobile liquid phase, ~ is a first-order mass-transfer coefficient describing the rate of transfer between the mobile and immobile liquid phases (T- ~ ), and ~21m , flsm and fllim, flsim are the first-order degradation coefficients for the liquid and sorbed phase in the mobile and immobile regions (T- ~ ), respectively. The assumptions underlying the mobile- immobile concept are discussed extensively in the literature (Van Genuchten and Wierenga, 1976; Nkedi-Kizza et al., 1983; Goltz and Roberts, 1986).

While at least in theory one can define four different degradation coefficients, as seen above, in practice it is nearly impossible to distinguish between these degradation processes and to find suitable parameter estimates. Thus, the number of degradation terms must be reduced. One simplification is the

TRANSPORT OF PHTHALATE-ESTERS 1 15

assumpt ion that all rate coefficients are the same, i.e.,/~tm = ~sm = ] ' / | i m =

/A~m = /~* (Van Genuchten and Wagenet , 1989). With this simplification eq. 6 and 7 can be expressed in the following nondimensional form:

fiR t3ct _ 1 c32ct t~cl co(cl -- c2) -- c l f l R ~ (8) 0T P Ox 2 Ox

?c2 = co(cl - c2) - c2(1 - / ~ ) R ~ (9) (1 - fl)R ~ T

by defining the following dimensionless parameters:

T - VtL -- Vmt~'L ' To - Vt°L (10)

2 x = Z ( 1 1 )

m ~b = --0-; 0 = 0~ + 0~m (12)

p _ v,nL = v L for Om D +

0 m + pfKp ~ = o + p z~p

ctL ~L

Vm Om q

la*L ¢ , = 7.)

Dm0m D + - (13)

0

(14)

( 1 5 )

(16)

C m c, - (17)

C O

Cim c2 = - - (18)

CO

where L is an arbitrary positive distance f rom the origin. Analytical solutions to eq. 8 and eq. 9 for the following initial and

boundary condit ions are given by Van Genuchten and Wagenet (1989):

e l (x , O) = c2(x, 0) = 0 (19a)

de1 (oo, T) = 0 (19b) Ox

116 T. ZURMOHL ET AL.

10 for 0 < T~< T 0 Cl(0, T) = (19c)

for T > T o

Equations 8 and 9 will hereafter be referred to as the physical nonequilibrium (PNE) model.

Model parameter estimation

The applicability of the described models to simulate the experimental data was evaluated using the curve-fitting computer program CXT4 of Van Genuchten and Wagenet (unpublished). The program is an extended version of CXTFIT (Parker and Van Genuchten, 1984b), including degradation in the PNE-model. The LEA-model contains the three independent parameters P, R, and ~k, whereas the PNE-model contains the five independent parameters P, R, r , 09 and @. The best way to evaluate the applicability of a model is to obtain all model-parameters by independent direct measurements and then to compare the predicted BTC's with experimental values. But direct determi- nations of parameters like D or fl are possible only if the geometry of the porous medium, i.e., aggregate shape and size, is known. As we used undisturbed soil columns, fitting of the analytical solutions of the two models to the observed effluent data was necessary in order to obtain estimates of the transport parameters. Nevertheless, the number of parameters optimized simultaneously should be kept as small as possible.

M A T E R I A L A N D M E T H O D S

Soil

The soil used in the experiments was a Typic Haplaquept derived from mesozoic sandstone sediments. The texture is loamy sand. The samples were taken from a field location near the town of Bayreuth in Bavaria, Germany. The site has been in agricultural use for some decades, but had before been covered with a forest. Accordingly, the top 30 to 35cm of the soil is a homogeneous, humic, slightly aggregated Ap-horizon. The subsoil (Bv- horizon) is homogeneous as well, but less humic. There are occasional macropores from decayed tree roots with higher organic carbon content. Due to a clay layer at about 3.5 m depth there is a relatively high ground water level in the winter and spring (up to 40 cm below surface).

The soil samples selected for the column experiments were taken from the Ap-horizon and from the Bv-horizon. The major physico-chemical properties of the soils are given in Table 1.

TRANSPORT OF PHTHALATE-ESTERS

TABLE 1

Soil characteristics

117

Soil Depth pH org.C (fo~) Sand Silt Clay (cm) (0.01 M CaCI2) (%) (%) (%) (%)

Ap 0-35 4.8 0.8 77.1 19.0 2.9 Bv 35-60 3.9 0.1 72.5 19.1 8.3

TABLE 2

Physico-chemical properties of phthalate esters

Molecular Aqueous log Kow Vapour weight solubility pressure (gmol -I) (mgi -~) (25°C) (Pa, 25°C)

DMP 194.2 4290 1.53-2.1 < 0.001 DEP 222.2 928 2.35 0.37 DBP 278.4 10.1 4.57 0.004 BBP 312.4 2.8 4.2-5.0 0.002 DEHP 390.6 0.04-0.4 5.3-8.7 4" 10 -5

Chemicals

Five phthalte esters (PE's) included in the Environmental Protection Agency's list of priority pollutants were selected for the experiments: dimethyl phthalate (DMP), diethyl phthalate (DEP), dibutyl phthalate (DBP), butyl- benzyl phthalate (BBP) and di-(2-ethylhexyl) phthalate (DEHP). Table 2 summarizes the physico-chemical properties of these PE's.

Experimental devices

Breakthrough characteristics were determined for undisturbed soil cores (9.4 cm diameter, 14.5 cm long), kept in stainless steel cylinders. The usual way to get small undisturbed cores is to drive a sharpened cylinder into the soil by hammer blows with the help of a piston. Preliminary experiments showed that driving the larger cylinders into the soil without any guideway for the piston caused visible breaks and gaps in the soil core. Thus, an apparatus was used, where the piston is guided through two ball-bearings, which are connected with a sturdy strut. This design prevents any tilting of the cylinder and therefore yields soil cores without breaks or gaps between cylinder and soil core.

The cores were embedded in two brass endplates. Two chemically-inert porous plates (glassinter, air entry at - 7 kPa , ROBU, Germany) were put

1 1 8 T. Z U R M U H L ET AL.

,: / ) / / " ,g ts_

t

3c

3o

3e

5

A

2

Fig. 1. Experimental design. 1: Leaching solution; 2: Syringe pump; 2a: Syringe; 2b: Piston guide; 2c: Piston; 2d: Drive disk; 2e: Check valve; 3: Soil monolith; 3a: Steel cylinder; 3b: Porous plate; 3c: Brass endplate; 4: Effluent sample; 5: Stop cock; 6: Hanging water column device.

between the soil and the lower and upper endplate. Two rubber O-rings made the bottom plate of the column assembly air-tight.

The leaching solution was applied to the top of the column with a multi- channel syringe pump (Soil Measurement Systems, Tucson, AZ, U.S.A.). The porous plate at the inlet ensured that the solution was evenly spread over the soil surface.

To keep the soil cores unsaturated the solution was applied at a flux-rate smaller than the saturated hydraulic conductivity of the soil. Suction was applied at the outlet of the columns. This suction was produced and regulated with a hanging water column of length H. The whole experimental design is depicted in Fig. 1. Bulk density, water content and degree of saturation were determined individually for each soil column after termination of the leaching experiments. The adjusted flux and the applied suction lead to steady state flow conditions with an essentially uniform water content throughout the soil columns.

Analytical methods

The phthalate esters were extracted from the collected water fractions with hexane and analyzed by a gas chromatograph equipped with a fused-silica capillary column (DB-5) and an electron capture detector (ECD) (Russell et al., 1985). Recovery studies with spiked soil-water solutions were carried out


TABLE 3a

Experimental conditions

119

Exp Compound soil q 0 0/e p v T O (cmh -l) (%) (%) (gcm 3) (cmh-~) ( _ )

Apl C1 A 0.21 33.7 82.8 1.54 0.61 0.53 Ap2 PE's A 0.21 33.7 82.8 1.54 0.61 1.87 Bvl C1 B 0.35 28.6 80.1 1.69 1.22 0.61 Bv2 PE's B 0.35 28.6 80.1 1.69 1.22 2.13

e: porosity

TABLE 3b

Concentration of the input solutions

DMP DEP DBP BBP DEHP CI (/tgl -~) (/~g1-1) (/tgl -~) ( / tgl- ' ) (/~gl ~) M

6120 7190 3070 730 210 0.05

to quantify the extraction efficiency. Chloride was measured with an ion-selec- tive solid-state membrane electrode (INGOLD).

Experiments

BTC's for all five PE's were determined on one soil column each for soil Ap (Exp. Ap2) and Bv (Exp. Bv2), respectively. After the termination of the PE-percolation experiments, BTC's of chloride were measured on the same soil columns and at the same flux rates as for the PE-experiments (Exp. Ap 1 and Bvl, respectively). Tap water was used as leaching solution with 0.05% NaN 3 (w/w) added to avoid microbiological degradation of the PE's, as described by Russell and McDuffie (1986). The tap water had an ionic strength of about 0.008 M, with Ca being the major cation species. The ionic strength therefore lies between that of rain water, which is "applied" under natural conditions, and that of a soil stabilizing 0.01 MCaCl-solution, as it is recommended for percolation experiments in the literature.

All experiments were run at ambient temperatures of 22-24°C. A summary of the experimental conditions for all experiments is given in Tables 3a and 3b.

120 T. ZURMIDHL ET AL.

a LEJ~-Mode l ....~.~ I~. ~..o~ ~1,._. O - 1 . 6 8 cm~/h O*- 1.81 cm~/h

0.6 R - 1.00 R - 1 .00 p - 0 . 0 0 h -t O " 1.00

- 999 .66 05 - 0.00

~0 0,4 •e

~ 0.3

t~ 0.1 ,,,, e • •

0.0 - -~ • - ~ 1 0.0 10 2.0 3.0 4.0 ~.0

P o r e V o l u m e T

b LEA-Model . .. ~N.]~.-:.bl~:l~.L. O -2 .10 cma/h 0 % 2 .05 c ~ l h 0.6- R - 121 R - 1 .20

, ~ u ° 0 . 0 0 h- ' B - 0 .92 0.5- ~ - 1.45

=o o,- - "~ ~ o.3-

~ 0.2-

0.0 0o

Pore Volume T Lo

Fig. 2. Measured and calculated BTC's for C1- on soil-type Bv (Exp. Bvl), (a) R set to unity, (b) R allowed to vary,

RESULTS A N D DISCUSSION

Tracer breakthrough curves

Figures 2 and 3 show the measured chloride BTC's of the experiments Bvl and Ap 1. Also shown are the fitted BTC's using either the LEA-model or the PNE-model. In each figure the experimental BTC's are presented as dots and the calculated BTC's are shown as solid lines. The parameters, which were actually used in the optimization procedures, are given in the upper part of the figures.

Using the LEA-model, only the dispersion coefficient, D, was optimized. For the PNE-model the Peclet number, P, and the parameters fl and to, which indicate the extent of nonequilibrium, and the kinetics of reaching the equilibrium, respectively, were subject to the fitting procedure. The pulse pore volume, T 0, and the Darcy-flux, q, were measured during all experiments. The average pore-water velocity, v, was then calculated using the relationship v = q 0-1. The volumetric water content, 0, was determined from weight

t.F~-Modea ....E.~J;.v.M~gL.. O - 3.74 cma/h O % 2.36 cmJ/h

0'6 7 R - l . O 0 R - 1 .00 | A ~ -0.00 h-' B - 0.55 I I . ~ \ . - 2 . 2 5 ,°' 1 . o o o

~ 0.2 ",,,,

0 .0 ,

0.0 0.5 1.0 1.5 2.0 2.5 3.0 P o r e V o l u m e T

Fig. 3. Measured and calculated BTC's for C1- on soil-type Ap (Exp. Apl).


measurements after termination of the experiment. For both models the degradation coefficient was set to zero and R was set to unity.

Figure 2a (Exp. Bvl) shows that both models failed to describe the experimental data of the chloride transport in soil Bv adequately. This is caused by a shift of the measured chloride peak to a pore volume position larger than one, indicating anion adsorption. This may be explained by an ion-exchange process. After the percolation experiments with the PE's, where tap water with low ion content was used, most of the ion exchange sites were occupied with H+-ions. By changing the input solution to 0.05 MKCI in order to measure the chloride-BTC's, these H+-ions were replaced by K+-ions, which sub- sequently caused a decrease of the soil solution's pH to values below 3.5. This low pH-value then caused the pH-dependent charge of inorganic soil components, i.e., oxides and hydrous oxides of Fe, AI and Si, to become positive, in turn making anion adsorption possible (Espinoza et al., 1975).

Because of the anion adsorption in soil By, the fitting procedure was repeated with R left variable. Both models yield almost identical curves (Fig. 2b) with the same value of R and similar estimates of the dimensionless parameters P (Tables 4 and 5). The value of ~ is close to unity. Thus, physical equilibrium is established in this soil column, and the LEA-model is appro- priate for describing the physical transport phenomena.

For the more aggregated soil Ap (Fig. 3) the PNE-model yields a better fit to the observed data than the LEA-model. This indicates that under the given experimental conditions not all of the wetted pore space actively takes part in the transport process, i.e., a physical nonequilibrium is established. The assumption that R is equal to unity is justified for soil Ap, because fitting the data with the PNE-model with R left variable yields an R value of 1.03. No anion adsorption is observed in the Ap, probably because the Ap has a lower content of iron oxides and a higher H + buffer-capacity.

Assuming a linear relationship between the dispersion coefficient D(Dm) and the dispersivity 2 as D = 2 • v (D m = 2 • v~) (Nielsen et al., 1986), values for 2 can be determined (Tables 4,5). Comparison with literature values shows that the values for 2 reported here are higher than those reported for disturbed soil columns (0.01 cm < 2 <0.5cm) (Nkedi-Kizza et al., 1983; Wierenga, 1988; Khan and Jury, 1990), but are in good agreement with experiments on undisturbed soil columns (0.5cm < 2 < 20cm), as reported by Smettem (1984), De Smedt et al. (1986), and Khan and Jury (1990).

The reported results of the CI--BTC's are typical for the two horizons, which is supported by a series of tracer experiments performed in order to characterize spatial variability (not shown here).

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1 2 4 T. ZURMOHL E T A L.

(~ LEA-Model ....~. ~:.-. _ _~..~l_~].... O - 2.10 c~lh 0 % 2,10 cm21h 1.0" • • R -1.45 R - 1.46

~ p -000 h-' . - 0.92

. - 2512.54 ~ o.B / ~ ~-o.oo ~ 0.6.

O.4.

i~ O2-

o.o oo ~.o ;.o do d.o ,6.o Pore Volume T

b I,O

R - 1,58 R -

" . ' B -

0

~ 0.6"

~ 0.4-

~ 02-

0,0 : ~ : i o.o 2.0 4.0 6.0 8o

P o r e Volume T

LEA-Model ._ E N.~ .-Mcglgl....

0 = 2.10 cm'/h 0 % 2.10 cm~/h 1.52 1.00 1200.00 0.00

10.0

C LEA-Model

O - 2.10 cm'/h (I.05- R - 7.66 p - 0Z7 h-'

~ ° ~

"~ 003 -

0.02 =

> o01 -

0011 ,- o.o ,~,o

... _P..NI~_T.bL .~. eJ... 0 % 2 .10 cm'/h R - 7.79

- 0.96

- 157.24 # - 0.42

~0 ,~o ,;o z6o Pore Voluffae T

d LEA-Model ._. P..N ~_ =.M. _ocLel__ ,.

O =2,10 cma/h 0 +- 2.10 cm~/~ 0.010 ,q - 12.B9 R - 13.55 ~ -047 h -~ B - 0.41

~ - 64.97 0.006 - - 0,41

o: 0006 =

o.o114-

0 002 -

0000 ,~ - , , --,

o.o ;o ..o ,zo , ,o ~.o Pore Volume T

Fig. 4. Measured and calculated BTC's for PE's on soil-type Bv (Exp. Bv2), (a) DMP, (b) DEP, (c) DBP, (d) BBP.

Breakthrough curves of phthalate esters-soil Bv

Figure 4a-d shows the measured and fitted BTC's of DMP, DEP, DBP, and BBP, respectively, for soil Bv (Exp. Bv2). No breakthrough of DEHP could be detected during the percolation of about 30 pore volumes. The different peak retardation times, which are in accordance with the peak maximum concentrations, show clearly that the mobility of the PE's decreases with increasing molecular weight, decreasing water solubility and increasing values of the octanol-water distribution coefficient /Cow (Table 2).

The observed PE BTC's were fitted by means of the LEA-model. To keep the number of fitting parameters as small as possible and to verify the model assumptions we used the dispersion coefficient obtained from the chloride BTC of Exp. Bvl for the fitting procedure. We cannot exclude that the different composition of the percolation solution in the CL- and the PE experiments had an effect on the dispersive properties of the soil, leading to a different D. However, our approach yields good results, as discussed below.


TABLE 6

Percent mass recoveries of phthalate esters

125

Soil % Mass recovery

DMP DEP DBP BBP DEHP

A 45.7 31.8 n.b. n.b. n.b. B 101.3 101.7 8.8 1.9 n.b.

n.b.: no breakthrough observed.

Further, if D and R are simultaneously optimized, they are often inter- correlated, thus leading to large confidence intervalls.

The degradation coefficient was set to zero. Hence, the only optimized parameter was R. The measured BTC's of DMP and DEP are well described by the model (Fig. 4a, b). Thus, the local equilibrium assumption is valid in a physical as well as in a chemical sense.

For the more hydrophobic DBP and BBP the LEA-model failed to describe the measured breakthrough on soil Bv. Fitting the PNE-model with varying R, fl and to brought no improvement. The values of the parameter fl are close to one, the value of co (which describes the rate at which equilibrium is obtained) is large (Table 5). The large 95% confidence intervals of these parameters indicate the uncertainty in finding a unique solution for the PNE-model.

In order to get a reasonable fit it was necessary to optimize the degradation coefficient in the fitting procedure, although microbiological degradation was minimized by adding NaN 3 to the percolation solution. With non-zero values of p the LEA-model describes the BTC's satisfactorily. An effluent mass balance was carried out for all PE-experiments by comparing the size of the input pulse to the area under the BTC's. The results (Table 6) indicate that only a small part of the input mass of DBP and BBP passed the soil column of soil Bv. The efficiency of NaN3 in preventing microbiological degradation is documented in the literature (Russell and McDuffie, 1986; Pecher et al., 1990). Further, a batch experiment with a soil solution spiked with PE's and 0.05% (w/w) NaN 3 added showed no decrease in PE concentrations over a period of two weeks. Degradation experiments for PE's in soils under optimal conditions for microflora were performed by Shanker et al. (1985), Inmann et al. (1984) and Schmitzer et al. (1988). Shanker et al. (1985) found that DMP and DBP were degradated to 10% and 8% of the initial concentrations, respectively, within 10d. Inmann et al. (1984), on the other hand, reported a lag-phase of 10 to 20d before DBP degradation was initiated. Schmitzer et al. (1988) gave half-life times for the degradation of

126 -r. ZURM~HL ET AL.

DEHP to be about 95d. Hence, even if the NaN 3 was not completely successful in preventing microbial degradation, we conclude that microbiological degradation cannot be the exclusive cause for the observed mass deficits.

Callahan et al. (1979) investigated abiotic degradation of PE's by hydrolysis. They reported half lives in water at pH 7 from 3.2 y for DMP to 2000 y for DEHP. They concluded from their findings that abiotic degradation does not play a significant role in the overall degradation process.

Losses by sorption to the experimental apparatus is thought to be negligible since it is made of chemically inert materials, and the exposed surface area of the apparatus is very small in comparison to the surface of the soil particles. Further, complete mass recovery for the DMP and DEP on soil Bv demon- strates that the mass balance error in soil Ap (discussed below) can not be due to sorption losses to the experimental device.

Thus, a part of the total PE-mass was still in the soil column due to chemical nonequilibrium and/or an irreversible adsorption. These findings are supported by Ogner and Schnitzer (1970), who showed that fulvic acids can complex phthalates, and that these PE's can not be re-extracted with organic solvents. Whereas the general fit of the DBP and BBP BTC's appears to be satisfactory, the measured data show, at the falling sections of the BTC's, a more pronounced tailing than the fitted curves. This again suggests that the observed mass deficit can not be caused by a degradation process alone. Thus a chemical nonequilibrium exists, which is caused by an apparent adsorption/desorption hysteresis from different adsorption/desorption kinetics due to a very slow desorption process.

Equations 3 and 4a show that under the assumption of local equilibrium it is impossible to trace back to what extent/~ is composed by a first order decay in either the solid phase (Ps) and/or the liquid phase (P0. Thus, assuming irreversible first order surface complexion (and no biological decay) we can calculate the rate coefficient by Ps = P (R-l) -t. Computed values of Us are given in Table 4.

Using the values of R and eq. 2 it is possible to determine Kp-values for the four PE's. The distribution coefficients for DBP and BBP can now be considered adsorption coefficients rather than real distribution coefficients, because the desorption is not in equilibrium with the adsorption, as discussed above. Considering the adsorption to take place at the organic carbon fraction foc in soil, the value of Kp may be transformed to the more universally valid distribution coefficient between water and organic carbon Ko¢ according to Kp = Koc" foc- Values of the Koc are given in Tables 4 and 5. These values are compared with experimental values from the literature, or, if no experimental data were found, are compared with calculated values using the empirical

T R A N S P O R T O F P H T H A L A T E - E S T E R S | 27

c l L E A - M o d e l

D = 3.74 cm~/h D*- 2 .36 cm~/b 0.40- R - 3 2 0 R - 3 . 7 0

u - 006 h-' # - 0 . 5 5 . = 0.44

0.32~ .,."'~ ~ - 0. '10 / '-

016-

OOB

0 O0 -

oo 20 o


... £NJ~.~. _~_.od..el .. b L E A - M o d e l _ ...~.E.s.. _M.ocLe!...

o~5- O - 3 7 4 c~/h 0 ÷- 2.36 cm~/b R - 4 . 6 0 R - q .60 V =0.08 h " t3 - 0 . 5 5 a

~ - 1.10 0~'O - ,,'Q, =o . / ,, ~ - 0 .40

010


I

120

Fig. 5. Measured and calculated BTC's for PE's on soil-type Ap (Exp. Ap2), (a) DMP, (b) DEP.

relationship between log Kow and log Koc, as given by Schwarzenbach and Westall (1981):

log Koc = 0.72 log Kow + 0.49 (20)

The agreement between the values obtained from the BTC's and the values from the literature is reasonably good. Thus, using the dispersion coefficient from the C1- BTC and Kp values from literature would have lead to a successful prediction of the measured BTC's for DMP and DEP.

Breakthrough curves o f phthalate esters - soil Ap

Figure 5a,b shows the measured and calculated BTC's for DMP and DEP for soil Ap. No breakthrough could be measured after percolation of about 30 pore volumes for DEHP and, in contrast to soil Bv, for DBP and BBP. The difference in the mobility of the PE's in the soils Ap and Bv can be attributed to the higher organic carbon content of soil Ap, as hydrophobic sorption is the predominant sorption mechanism for hydrophobic organic compounds like phthalate esters. Because there was no breakthrough of DBP, BBP and DEHP, the discussion is limited to DMP and DEP only.

As for soil Bv, the dispersive properties from the C1- BTC were taken for the simulation of the PE-BTC's. In order to keep the number of fitting parameters as small as possible, the retardation coefficients were also determined independently by using the values of Koc obtained from the BTC's on soil Bv, the foe-value of soil Ap and eq. 2. The Koc values appear to be reliable because they result from experiments with a complete mass recovery and because they are in accordance to values from literature.

Assuming that the sorption sites and the mobile water fractions are dis- tributed in the same manner throughout the soil system (Nkedi-Kizza et al.,

128 ~. Z~;RMOHL ET AL.

1984; Brusseau et al., 1989) one can simplify eq. 14 to

0m /3 = f = --0- = q~ (21)

This means that/3 represents the portion of the water in the soil column which is mobile. This portion should not change between two different solutes if the flow domain is kept constant. Thus, the parameter/3 in the PNE-model was also fixed to the value obtained with the C1- BTC. In contrast to soil Bv, the mass recovery of DMP and DEP at the outlet of the soil column was not complete (Table 6). Thus, only the coefficients ~O and co for the PNE-model and ~, in the LEA-model were variable optimization parameters.

From Fig. 5a,b it is obvious that the LEA-model in this case fails to describe the measured BTC's. It cannot adequately describe the early breakthrough and the high solute peak. The PNE-model, however, describes the BTC quite well, although only two parameters were left variable. The failure of the LEA-model and the agreement of the PNE-model to describe the measured BTC's of the two PE's is in consistency with the results of the breakthrough of chloride on the same soil column (Exp. Apl), which already suggested a physical nonequilibrium in soil Ap.

The incomplete recovery of the PE's and the thereby resulting requirement of a non-zero sink term again indicate chemical nonequilibrium, as was discussed above for soil Bv. As for DBP and BBP on soil Bv (Fig. 4c, d) the measured BTC's still show a more pronounced tailing than the simulated curves. This indicates that in the more humic soil also the less hydrophobic PE's are subjected to chemical nonequilibrium.

We want to stress that an apparently successful curve fitting for the data of Figure 5a, b can be achieved with the PNE-model, even if the degradation coefficient is set to zero. This would lead to a retardation coefficient which is forced to be 30 times higher than the one estimated by considering degradation. It is also possible to achieve a better fit of the measured BTC's of DMP and DEP with the LEA-model if the retardation coefficient is allowed to vary. This, however, yields values for R which are 2-3 times smaller than those predicted independently from soil Bv and which have shown to be in good agreement with other investigations. Thus, particular care should be taken when estimating several parameters simultaneously from curve fitting procedures.

General discussion

Combining the results of the different experiments we come to the following findings: - Under the applied flow conditions, in soil-type Bv a physical equilibrium is


established. Using the PNE-model yields no improved description of any of the measured BTC's. A chemical nonequilibrium exists for the phthalate

. esters with a higher molcular weight. This nonequilibrium increases with decreasing solubility of the PE's, as indicated by a higher apparent degradation coefficient.

- In soil Ap, a physical nonequilibrium exists as indicated by the CI- BTC. A chemical nonequilibrium ensues already for the less hydrophobic PE's. Thus, the degree of chemical nonequilibrium increases with increasing organic carbon content of the sorbent and with decreasing solubility of the solute. For sorbing solutes, sorption isotherm nonsingularity, i.e., sorption/

desorption hysteresis, and sorption kinetics can contribute to chemical nonequilibrium during solute transport. Further, asymmetric BTC's can be due to sorption isotherm nonlinearity. Sorption isotherm linearity appears to be a reasonable assumption for low solute concentrations. As demonstrated by Chiou et al. (1979) and Munz and Roberts (1986) linear isotherms may be expected for solution concentrations below 0.056 M, or close to the hydrophobic solute's solubility, whichever is lower. Thus, sorption linearity can be assumed for the PE's in this study. This assumption is further supported by an investigation of Russell and McDuffy (1986), who reported Kp-values for measured PE-sorption isotherms.

Chemical nonequilibrium is often thought to arise by two sorption sites of different sorption mechanisms (Cameron and Klute, .1977). Sorption is assumed to be instantaneous for one site and rate-limited for the other. Nkedi-Kizza et al. (1984), however, demonstrated that a two-site-model with these assumptions is mathematically equivalent to the PNE-model with respect to the derived BTC's. Thus, no differentiation of the processes taking place in soil columns is obtained with this model. Further, there is support in the literature showing organic matter to be the predominant sorbent for hydrophobic organic compounds, with mineral-surface contributions to sorption being negligible (Brusseau and Rao, 1989). This suggests that generally the assumption of two different sorption sites, as required by the two-site-model, may not be viable. Therefore, we propose that the observed chemical nonequilibrium phenomena are due to slow nonaqueous phase diffusion into the organic matter (IOMD) and subsequent slow mass-transfer between the liquid and solid phase and/or subsequent sorbate entrapment, i.e., irreversible adsorption. Similar results are reported by Bouchard et al. (1988) and Nkedi-Kizza et al. (1989), who attributed observed chemical nonequilibrium to IOMD. However, since they did not perform desorption experiments, they could not investigate possible irreversible adsorption. Note that both nonequilibrium mechanisms, i.e., irreversible adsorption and IOMD can occur on the same soil sample, as pointed out by Brusseau and


Rao (1989). The sorption hysteresis or irreversible adsorption may be considered as the limiting case of IOMD-controlled desorption, when the solute is entrapped inside the organic matter, so that desorption is no longer possible. Findings from batch experiments support this theory for chlorophenols. Isaacson and Frink (1984) reported that desorption was slower than sorption and that in some cases up to 90% of the sorbate was irreversibly held in the soil. Evidence for IOMD-limiting mass transfer par- ticularly for PE's is given by a study of Freeman and Cheung (1981). They showed (a) that elution of DEHP from a contaminated sediment with flowing water was unsuccessful due to irreversible adsorption and (b) that DEHP desorption was most complete with eluting solvents causing maximum swelling of the soil organic carbon matrix and therefore minimizing the diffusive mass transfer.

Our findings support the conclusion of Brusseau and Rao (1989) that physical nonequilibrium and sorbate diffusion within the matrix of sorbent organic matter (i.e., IOMD) are the predominant rate limiting mechanisms for the sorption of hydrophobic organic compounds. More attention, however, should be given to the process of entrapment of hydrophobic organic compounds in the soil organic matter leading to incomplete release of these compounds from soil.

SUMMARY AND CONCLUSIONS

Breakthrough curves of chloride as a tracer and phthalate esters on undisturbed and unsaturated soil columns during steady water flow conditions showed that the mobility of the phthalate esters decreased with increasing Kow-values and increasing organic-carbon content of the sorbate. Calculated Koc values were in accordance with values from batch experiments reported in the literature.

The movement of chloride in the subsoil was well described by the LEA- model, showing physical equilibrium. The tracer-BTC in the slightly aggregated topsoil with a higher organic carbon content required the PNE- model for a proper description, indicating the existence of pore domains accessible by diffusion only, and hence a physical nonequilibrium.

For both soils, chemical nonequilibrium was observed with the phthalate esters. The chemical nonequilibrium, indicated by an incomplete mass balance and a very long tailing at low concentration, increases with decreasing solubility of the PE's and with increasing organic carbon content of the soil. The phenomenon is attributed to a slow diffusion of the contaminants into the organic matrix of the soil and a possible subsequent entrapment of these substances in the organic matter.

When modelling the PE BTC's it was possible to account for the chemical


nonequilibrium using the physical parameters from the tracer BTC only if additionally a non-zero sink term was included. Since the observed mass deficits of the PE's cannot be completely due to microbiological degradation, it is concluded that both models fail to describe the observed nonequilibrium phenomenon. Optimization of PE BTC's with the PNE-model with zero degradation but with all other model parameters allowed to vary, lead to an apparently satisfactory fit, but resulted in senseless parameter values.

Our results show that transport of PE's is influenced not only either by a physical or a chemical nonequilibrium but that both nonequilibrium processes appear simultaneously. Brusseau et al. (1989) already presented a model that accounts for multiple sources of nonequilibrium. Irreversible adsorption, however, was not included. Hence, further research is needed for the investigation of IOMD and irreversible adsorption of organic substances in soils and for the account of these phenomena in deterministic solute transport models.

In assessing the mobility of PE's in a field soil, the spatial and temporal variability of all parameters, as well as their possible functional dependence on water content, flux rates etc. has to be taken into account. Currently we are investigating some of these aspects for different soils in a series of laboratory experiments. Even more severe problems lie in the attempt to come to an universally valid description of the degradation processes of phthalates in a natural environment.

Apart from the parameter estimation problems mentioned above there remain fundamental problems in applying results from laboratory experiments to the field (Nielsen et al., 1986). Important examples are the validity of the underlying flow model, and the apparent scale dependency of the dispersion coefficient.

ACKNOWLEDGMENTS

This study is part of a project funded by the Federal Environmental Protection Agency of Germany (Umweltbundesamt) under contract number 107 01 016/03.

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transport of phthalate-esters in undisturbed and unsaturated soil columns

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