Fate of seven pesticides in an aerobic aquifer studied in columnexperiments

Nina Tuxen *, Peter L. T�uchsen, Kirsten R�ugge, Hans-Jùrgen Albrechtsen,Poul L. Bjerg

Department of Environmental Science and Engineering, Groundwater Research Centre, Technical University of Denmark, Building 115,

DK-2800 Lyngby, Denmark

Received 5 July 1999; accepted 21 October 1999

Abstract

The fate of selected pesticides (bentazone, isoproturon, DNOC, MCPP, dichlorprop and 2,4-DD) and a metabolite

(2,6-dichlorobenzamide (BAM)) was investigated under aerobic conditions in column experiments using aquifer ma-

terial and low concentrations of pesticides (approximately 25 lg/l). A solute transport model accounting for kinetic

sorption and degradation was used to estimate sorption and degradation parameters. Isoproturon and DNOC were

signi®cantly retarded by sorption, whereas the retardation of the phenoxy acids (MCPP, 2,4-DD and dichlorprop), BAM

and bentazone was very low. After lag periods of 16±33 days for the phenoxy acids and 80 days for DNOC, these

pesticides were degraded quickly with 0.-order rate constants of 1.3±2.6 lg/l/day. None of the most probable degra-

dation products were detected. Ó 2000 Elsevier Science Ltd. All rights reserved.

Keywords: Pesticides; Herbicides; Aerobic degradation; Kinetic sorption; Aquifer; Modelling; Column

1. Introduction

The detection of pesticides in groundwater in the

recent years indicates that application of pesticides has

caused these substances to spread in the environment

(Barbash and Resek, 1996). In Denmark pesticides were

found in 17% of 4209 drinking water wells analysed

during the period 1989±1997 (GEUS, 1998). More than

99% of the drinking water in Denmark is obtained from

groundwater which undergoes only simple treatment,

such as aeration and ®ltration. Consequently the nu-

merous ®ndings of pesticides in groundwater constitute

a severe threat to the drinking water supply in Denmark.

The present investigation focuses on seven pesticides,

which are some of the most frequently found pesticides

in Danish groundwater. The selected pesticides are all

herbicides and are moderately soluble and include both

non-polar compounds and dissociable compounds

(Table 1).

Most of the reported investigations of pesticide de-

gradation and sorption have been conducted using

topsoil and concentrations in the mg/l-range. Very lim-

ited knowledge is available regarding the fate of pesti-

cides in aquifers, characterised by low contents of

organic carbon, low density of microorganisms and low

concentrations (lg/l) of the pesticides. However, a few

investigations concerning the phenoxy acids have been

reported (Agertved et al., 1992; Heron and Christensen,

1992; Klint et al., 1993) who observed relatively high

degradation rates (2±6 lg/l/day) for MCPP under aero-

bic conditions in sandy aquifers after lag periods of

about 50 days. Also 2,4-DD has been degraded under

Chemosphere 41 (2000) 1485±1494

* Corresponding author. Tel.: +45-45-25-15-95; fax: +45-45-

93-28-50.

E-mail address: nit@imt.dtu.dk (N. Tuxen).

0045-6535/00/$ - see front matter Ó 2000 Elsevier Science Ltd. All rights reserved.

PII: S 0 0 4 5 - 6 5 3 5 ( 9 9 ) 0 0 5 3 3 - 0

aerobic conditions using aquifer material in laboratory

batch experiments (Kuhlmann et al., 1995). These

studies revealed that sorption of the phenoxy acids was

very limited in sandy aquifer material. Aerobic degra-

dation and a small, but signi®cant, sorption of isopro-

turon was found in laboratory batch experiments with

chalk aquifer materials (Johnson et al., 1998).

The literature is very sparse regarding the fate of

DNOC, BAM and bentazone in groundwater and other

environments. Nitroaromatic compounds (including

DNOC) can sorb strongly to some speci®c clay minerals

indicating that sorption of DNOC could be important

even in aquifers with low organic carbon contents

(Haderlein et al., 1996; Weismahr et al., 1997). Degra-

dation of bentazone was found in soil with a high or-

ganic carbon content (Wagner et al., 1996), but to our

knowledge no investigations of these compounds using

aquifer materials have been reported.

Most of the above-mentioned results have been ob-

tained by laboratory batch studies where the degrada-

tion and sorption properties are investigated separately.

Recent investigations indicate that batch experiment

may not be appropriate when determining degradation

rates since degradation rates can be dependent on the

Table 1

Physical and chemical properties of the pesticides

Chemical name Formula and

molecular

weight

(g/mol)

Structure formula Solubility

(mg/l)

log Kow pKa

MCPPa ± 2-(4-chloro-2-

methylphenoxy) propa-

noic acid

C10H11ClO3

214.6

734 (25°C) 1.26 3.78

Dichlorpropa ± ���-2-

(2,4-dichlorophenoxy)

propanoic acid

C9H8Cl2O3

235.1

350 (20°C) 1.77 3.00

2,4-DDa ± 4-dichlorophen-

oxy) acetic acid

C8H6Cl2O3

221.0

311 (25°C.

pH 1)

2.6±2.8 2.64

DNOCa ± 2-methyl-4,

6-dinitrophenol

C7H6N2O5

198.1

130 (15°C) 2.12b 4.31b

Isoproturona ± N,N-dim-

ethyl-N0-[4-(1-methyleth-

yl)phenyl] urea

C12H18N2O

206.3

65 (22°C) 2.5 ±c

Bentazonea ± 3-(1-meth-

ylethyl)-1H-2,1,3-enzo-

thiadiazin-4(3H)-one-

2,2-dioxide

C10H12N2O3S

240.3

570 (22°C) 0.77 (pH 5) 3.3

)0.46 (pH 7)

BAMd ± 2,6-dichlorben-

zamide

C7H5Cl2NO

201.0

nfe 1.54 ±

a Tomlin (1994).b Schwarzenbach et al. (1988).c Not relevant.d Verschueren (1996).e nf: Not found.

1486 N. Tuxen et al. / Chemosphere 41 (2000) 1485±1494

pore water velocity (Langner et al., 1998). A laboratory

column experiment is likely to be a better approach

because it is a ¯ow-through system which simulates the

interaction between sorption and degradation under

conditions comparable to actual aquifer environments

e.g. realistic solid/water ratios (Kookana et al., 1993;

Langner et al., 1998).

The purpose of the current study was: (a) to inves-

tigate the simultaneous sorption and degradation of

selected pesticides under ÔnaturalÕ aerobic, oligotrophic

conditions in columns using various combinations of

bentazone, isoproturon, DNOC, MCPP, dichlorprop,

2,4-DD and BAM (b) to investigate the di�erences in

pesticide behaviour between two sediments and (c) to

investigate the di�erences in degradation properties be-

tween the combinations of pesticides. A reactive solute

transport model accounting for kinetic sorption and

degradation was used to interpret the results obtained

(for example degradation rates, lag periods and Kd-

values).

2. Materials and methods

2.1. The Vejen aquifer

Groundwater and sediment for the column studies

were obtained from a shallow, uncon®ned, aerobic

aquifer located near Vejen, Denmark. The upper aquifer

is a glacio¯uvial sand and gravel aquifer (Weichselian,

Quaternary period). The thickness of the upper aquifer

is approximately 10 m and the water table is located 4±5

m below the ground surface. The geology, hydrogeology

and groundwater chemistry of this aquifer are described

in detail by Bjerg et al. (1992), Bjerg and Christensen

(1992) and Pedersen et al. (1991). The sediment which is

a medium to coarse grained sand, was collected 1 m

below the groundwater table at two di�erent locations

(location I and II) located 30 m apart. Groundwater was

collected near location II from the same depth as the

sediment. The organic carbon content of the sediments

was low (0.02%) and sediment II had a slightly higher

content of silt and clay than sediment I (Table 2). The

clay fraction of the sediment in the Vejen aquifer con-

sisted mainly of smectite, illite, kaolinite, and vermiculite

(Madsen et al., 1999).

2.2. The pesticides

The experiments were conducted using three di�erent

aqueous solutions of pesticides: solution A contained

2,4-DD (98.5% pure from Dr. Ehrenstorfer GmbH) and

MCPP (99% pure from Ridel-de-H�aen); solution B

contained BAM (97% pure from Aldrich), bentazone

(98% pure from Ridel-de-H�aen), DNOC (99% pure),

MCPP, dichlorprop (99% pure from Riedel-de-H�aen)

and isoproturon (99% pure from Ridel-de-H�aen); solu-

tion C contained isoproturon alone. Bromide (as NaBr)

was added as a conservative tracer to all the solutions.

The inlet concentrations were approximately 25 lg/l for

each pesticide. The bromide inlet concentration was

approximately 60 mg/l and did not signi®cantly in¯u-

ence the ionic strength of the groundwater.

2.3. Column experiment

The experimental set-up consisted of six stainless

steel columns each with a height of 105 cm and a di-

ameter of 10 cm (Fig. 1) packed with sediment under

saturated conditions. A stainless steel screen and a 2.5

cm thick layer of oven-dried, acid-washed gravel was

packed into both the inlet and outlet of the columns to

ensure a one-dimensional ¯ow and to prevent clogging

at the outlet. A peristaltic pump ensured a continuous

upward ¯ow of the pesticide solutions, thus maintaining

completely saturated conditions throughout the experi-

ment. The columns were kept at 10°C corresponding to

the average groundwater temperature in Denmark and

the pesticide solutions were placed in refrigerators at 0±

4°C to minimise the risk of degradation in the feed so-

lutions. Two di�erent sediments were used (I and II)

with three di�erent pesticide solutions (A, B and C) thus

a total of six column experiments were conducted (re-

ferred to as AI, AII, BI, BII, CI and CII).

E�uent samples were collected in 100 ml PE bottles

by a timed sample-collector during the experimental

Table 2

Properties of the aquifer materials

Parameter Unit Sediment I Sediment II

Organic carbona % 0.02 (0.006) 0.02 (0.005)

Gravel (2±4 mm) % 2.10 13.03

Sand (0.063±2 mm) % 97.31 85.66

Silt + Clay (<0.063 mm) % 0.59 1.31

Microbial density (plate counting using R2A agar)a;b 104 colony forming

units/g wet sediment

106 (28) 58 (12)

a The brackets contain the 95% con®dence interval.b Samples analysed immediately before the packing of the columns.

N. Tuxen et al. / Chemosphere 41 (2000) 1485±1494 1487

period of 140 days. The samples were ®ltered through a

45 lm polypropylene GMF ®lter and then frozen to

)18°C. Potential sorption to the bottles was tested for

by alternating sample collecting in glass bottles and in

PE bottles over a period of four days. One-sided vari-

ance analysis showed that at a 99% signi®cant level there

was no di�erence between the pesticide concentrations in

the two types of bottles.

Groundwater without pesticides was recirculated

through the columns for four weeks prior to the exper-

iment to establish equilibrium between water and sedi-

ment. The purpose of this procedure was to prevent any

interference by geochemical reactions having no rele-

vance to the present study. Physical and chemical

conditions (e.g. ¯ow rate, pH, O2 and pesticide

concentration) were monitored four times during the

experiment.

2.4. Analytical procedures

The pesticides were analysed by HPLC on a Perkin-

Elmer LC 235 diode array detector with a Perkin-Elmer

Binary LC Pump at wavelengths 205 and 220 nm. A

mixture of an organic and an aqueous eluent was used.

The detection limits were 2±3 lg/l. Analysis for potential

metabolites were made using a GC-MS (Hewlet-Packard

GC 6890 coupled with an MS 5973) with a Restek

15 m� 0:25 mm Stabilwax-DA capillary column. Sam-

ples were extracted with a Solid Phase Micro Extraction

(SPME) ®bre followed by a desorption at 270°C directly

in the GC-injection port.

Bromide, sulphate and chloride were analysed on a

Dionex ion chromatograph DX120. An Ion Pac AS14 4

mm (10±32) column (P/N 46124) was used in combina-

tion with an anion suppressor (ASRII 4 mm, self

regenerating). The cations were measured by a Perkin-

Elmer 5000 Atomic Absorption Spectrophotometer

using standard conditions. Dissolved oxygen was mea-

sured by the Winkler titration method on small volumes

(5 ml). The pH was measured in a mini ¯ow-cell with a

Metrohm 691 pH meter. Tritiated water (3H2O) was

analysed by liquid scintillation counting on a Packard

TriCarb 2000 with HiSafe 3 scintillation liquid. DOC

was determined on a O. I. Model TOC-analyzer and

sediment-bound organic carbon was determined in a

LECOâ-oven after removal of inorganic carbon by

treatment with 6% H2SO4.

2.5. Modelling

A classical advection dispersion model combined

with source-sink terms to account for the sorption and

degradation processes was used to quantify the results

obtained (Parker and van Genuchten, 1984; Toride et al.,

1995; Bjerg et al., 1996). The one-dimensional version

of the advection dispersion equation can be written as

oCot� DL

o2Cox2ÿ vp

oCoxÿ qb

hoSotÿ E; �1�

where C is the concentration of the pesticide [M Lÿ3], x

the distance from the inlet [L], t the time [T], and vp is

the porewater velocity [L Tÿ1]. DL is the sum of the

longitudinal dispersion and di�usion [L2 Tÿ1], DL �aLvp � D�, where aL is the longitudinal dispersivity [L]

and D� is the molecular di�usion coe�cient [L2 Tÿ1].

Here qb denotes the bulk density [M Lÿ3] and h the

porosity [)]. E is the source-sink due to degradation that

can either be 0.-or 1.-order. The sorption is modelled

using a kinetic bicontinuum sorption approach, de-

scribed as

oS1

ot� F � Kd

oCot;

oS2

ot� a��1ÿ F � � Kd � C ÿ S2�;

�2�

where S1 is the sorbed concentration in the instanta-

neous controlled domain [M Mÿ1 aquifer material] and

S2 the sorbed concentration in the kinetically controlled

domain [M Mÿ1 aquifer material]. F is the fraction of

sorbent for which sorption is instantaneous [)], Kd the

equilibrium distribution constant [L3 Mÿ1] and a is the

sorption rate constant [Tÿ1]. The concentration of the

total amount of sorbed pesticide is given by

Fig. 1. A schematic representation of the experimental set-up

used for the column experiments.

1488 N. Tuxen et al. / Chemosphere 41 (2000) 1485±1494

S � S1 � S2: �3�

A combination of two computer programs was used:

CXTFIT2 (Toride et al., 1995) and FLOW (Bjerg et al.,

1996). Both programs have the potential for modelling

kinetic sorption, but while FLOW includes three types

of degradation kinetics (Monod, 0.- and 1.-order),

CXTFIT2 can only model 1.-order kinetics. Addition-

ally FLOW has included a lag period in the approach.

3. Results and discussion

3.1. Physico-chemical conditions during the experiment

Tracer experiments with tritiated water, conducted in

all columns before the pesticide experiments began,

showed no signi®cant physical non-equilibrium pro-

cesses in any of the columns. The ¯ow rate was constant

during the experiment resulting in speci®c pore water

velocities for each column between 5.4 and 6.1 cm/day.

These velocities resulted in residence times in the col-

umns of approximately 20 days for the non-retarded

compounds. The porosities ranged between 0.31 and

0.34 and the dispersivities ranged between 3.1 and

8.5 cm. In general, there was a very good model ®t for

the tracer breakthrough curves with r2 > 0.995 for all

columns.

The groundwater had an oxygen content of 11.5 mg/l,

a nitrate content of 4 mg/l, an organic carbon content

of 2.1 mg/l, and an alkalinity of 0.4 meq/l (Table 3). This

inlet composition remained constant during the experi-

ment. The pH increased by about 0.5 pH-units, proba-

bly due to the progressive intake of atmospheric air into

the pesticide solution reservoirs compensating for the

loss of solute. It was also con®rmed that no degradation

of the pesticides took place in the feed solutions by

measuring the pesticide concentrations several times

during the experiment.

Most of the chemical properties remained constant

throughout the columns, but the concentration of O2

decreased by approximately 2 mg/l and TOC increased

by approximately 1.5 mg/l. The most likely explanation

is that some organic material was released during the

collection and handling of the sediment and that part of

this were subsequently degraded in the columns.

Despite the pH-increase over time at the inlet, the

outlet pH was constant, although signi®cantly lower

than at the inlet (pH of 5.3 on average in columns with

sediment I and a pH of 5.8 on average in columns with

sediment II). Because the total alkalinity was low

(0.4 meq/l), the dominating bu�ering system in the sed-

iment was assumed to be cation exchange. This was

supported by a parallel decrease in Ca2� concentrations

over the columns corresponding to the observed pH

decrease (the increase of H�).

3.2. Breakthrough curves

The sorption and degradation parameters were esti-

mated by comparing the breakthrough curve of each

pesticide in each column with the corresponding bro-

mide breakthrough curve. BAM, bentazone and iso-

proturon were all recalcitrant during the experiment and

while neither BAM nor bentazone were sorbed, isopro-

turon sorbed signi®cantly (Fig. 2). The breakthrough of

isoproturon appeared almost simultaneously with bro-

mide in the outlet, but over time isoproturon was de-

layed relative to bromide as a result of a time-dependent

(kinetic) sorption. All of the phenoxy acids (MCPP, 2,4-

DD and dichlorprop) were only sorbed slightly (Fig. 3) but

degraded after lag periods of 20±30 days. The occur-

rence of a lag period implies that the degradation was

microbial. Degradation refers to disappearance of the

mother compound, not necessarily to complete miner-

alization of the pesticides to CO2 and water. The

breakthrough curve of DNOC (Fig. 4) is more complex.

During he lag period of 80 days the breakthrough curve

depicts only transport and kinetic sorption but after

degradation begins, three processes are involved: trans-

port, sorption and degradation. The use of a reactive

solute model is then necessary in order to separate and

quantify the processes.

3.3. Sorption

A fully kinetic one-site sorption model �F � 0� was

used to simulate the sorption for all pesticides. Table 4

lists the resulting sorption parameters for all the pesti-

cides. The sorption rate constants a, ranged between

0.13 and 0.78 dayÿ1 which is on the same order of

magnitude as other a-values determined at similar pore

Table 3

Properties of the groundwater in the inleta

Parameter Groundwater

pH 6.4

Total alkalinity 0.4

O2 11.5

Clÿ 18.9

SO2ÿ4 10.9

NOÿ3 -N 4.0

Ca2� 26.8

Mg2� 2.0

Na� 9.7

K� 4.0

Total organic carbon 2.1

Microbial density (AODC)b 102 (29)

a All units in mg/l, except TAL (meq/l) and AODC (103 cells/

ml).b The brackets contain the 95% con®dence interval.

N. Tuxen et al. / Chemosphere 41 (2000) 1485±1494 1489

water velocities (Brusseau et al., 1991; Kookana et al.,

1993). The Dahmk�ohler number x, expresses the ratio

of hydrodynamic residence time to sorption reaction

time and thus, can describe the degree of sorption non-

equilibrium in a ¯ow system. In the present investigation

the estimated xÕs ranged between 1.2 and 4.5 suggesting,

that the use of non-equilibrium models was reasonable

when estimating KdÕs (Brusseau et al., 1991). As an ex-

ample the Kd of DNOC in column BII (the one with the

highest x) was calculated using both an equilibrium

model and a kinetic sorption model. The resulting KdÕswere very similar and only deviated by approximately

5%. However, it is obvious from Fig. 4 that the best

model ®t was obtained with the kinetic model. For

simulations with lower x-values the di�erence in KdÕsbetween the equilibrium and the sorption models were

up to 20%.

Sorption, represented by the non-equilibrium sorp-

tion model, was necessary in order to obtain a good

simulation of the breakthrough curves of the pesticides.

This is observed in many other column studies with

similar compounds and ¯ow rates (Sabatini and Austin,

1990; Kookana et al., 1993; Selim and Ma, 1995; Fortin

et al., 1997).

Fig. 2. Observed and simulated breakthrough curves for BAM,

bentazone and isoproturon.Fig. 3. Observed and simulated breakthrough curves for

MCPP, 2,4-D and Dichlorprop.

1490 N. Tuxen et al. / Chemosphere 41 (2000) 1485±1494

3.4. Estimation of sorption

In aquifers were the organic carbon content is low,

sorption to mineral surfaces may be of equally impor-

tance as sorption to organic material. Several relation-

ships for the prediction of Kd-values based on Kow or

solubility have been suggested in the literature. A rela-

tion that is especially suited for aquifers with foc < 0:1%

and logKow < 3:7 has been proposed by Piwoni and

Banerjee (1989)

log Kd � 1:01 log Kow ÿ 3:46: �4�

The predicted Kd-values (using Eq. (4)) are compared

with the Kd-values observed in the column experiments

in Table 5. BAM is a compound with a low log Kow-

value hence it is not expected to sorb signi®cantly and

the predicted Kd-value is in good agreement with the

observed KdÕs.

Bentazone and the phenoxy acids are weak acids and

at the actual pH-values in the columns only a very small

fraction (<1%) of these compounds is in the neutral

form. According to the general expectation that the

anionic form of a dissociable compound does not sorb,

no sorption of these compounds was observed.

Isoproturon is a non-polar compound. The isopro-

turon sorption was greater in sediment II than in sedi-

ment I and the predicted value corresponded to the

observed. Batch studies of the sorption of isoproturon in

chalk aquifers gave Kd-values in the range of 0.02±0.5 l/

kg (Johnson et al., 1998). These higher values could be

due to the fact that the sediment used had a higher or-

ganic carbon content (foc � 0:1ÿ 0:4%) or due to the

di�erences in experimental approach. In batch experi-

ments the physical contact between the sediment and the

groundwater is more extensive and this can lead to an

overestimation of the sorption potential in the aquifer

(Kookana et al., 1993).

The high sorption of DNOC was unexpected since

DNOC is a weak acid. At the current pH (6.0) only 2%

Fig. 4. DNOC ± observed and simulated breakthrough curves.

Simulated curves obtained with 0.-order or 1.-order degrada-

tion models. The equilibrium curve is obtained with equilibrium

sorption and no degradation.

Table 4

Sorption and degradation results for the pesticides obtained by column studies in the aquifer material from Vejen, Denmarka

Partitioning

coe�cients, Kd (l/kg)

Sorption rate

constant, a (daysÿ1)

Modelling lag period

(days)

0.-order degradation rate

constant, k0 (lg/l/day)

Sediment I II I II I II I II

MCPP (A) 0 0 0 0 33 16 1.3 1.5

MCPP (B) 0.04 0.04 0.14 0.41 31 24 1.6 2.6

Dichlorprop (B) 0.04 0 0.78 0 31 21 1.6 2.0

2,4-DD (A) 0.11 0 0.24 0 28 15 2.0 1.6

DNOC (B) ±b 0.32 ± 0.13 ± 80 ± 2.1

Isoproturon (B) 0.06 0.15 0.43 0.15 ndc nd nd nd

Isoproturon (C) 0.08 0.13 0.29 0.23 nd nd nd nd

Bentazone (B) 0 0 0 0 nd nd nd nd

BAM (B) 0 0.08 0 0.18 nd nd nd nd

a A, B and C refer to the pesticide solutions.b No results obtained.c No degradation observed.

Table 5

Observed and predicted Kd-values (using Eq. (4)) for the aquifer

material (Vejen, Denmark) obtained by column studies

Observed Kd (l/kg) Predicted

Kd (l/kg)Sediment I II

BAM 0.00 0.08 0.01

Bentazone 0.00 0.00 �0.00

MCPP 0.00±0.04 0.00±0.04 �0.00

Dichlorprop 0.04 0.00 �0.00

2,4-DD 0.11 0.00 �0.00

Isoproturon 0.06±0.08 0.13±0.15 0.11

DNOC ±a 0.32 0.001

a No result obtained.

N. Tuxen et al. / Chemosphere 41 (2000) 1485±1494 1491

of the compound is present in its non-dissociated form.

However, speci®c and strong sorption of DNOC and

other nitroaromatic compounds to the clay minerals

kaolinite, illite and montmorillonite has been observed

when the exchangeable cations associated with the clay

minerals are weakly hydrated (K�, NH�4 etc.) (Hader-

lein et al., 1996; Weismahr et al., 1997). Thus, the

presence of kaolinite and illite (Madsen et al., 1999)

may cause speci®c sorption of DNOC, which could

explain the extremely high discrepancy between the

calculated Kd-value and the observed Kd-values (a fac-

tor of 300).

In general the Kow relationship predicted the actual

Kd-values well, but for compounds such as DNOC

which are a�ected by a very speci®c sorption mechanism

the estimation equations can cause very misleading

results.

3.5. Degradation

As previously mentioned BAM, bentazone and iso-

proturon were recalcitrant during the experimental pe-

riod of 140 days and only DNOC and the phenoxy acids

were degraded (Figs. 2±4). Selected samples were

screened by a GC±MS for the most probable degrada-

tion products: the corresponding chlorophenols from

the phenoxy acids (2,4-dichlorophenol and 4-chloro-2-

methylphenol) and the corresponding cresols from

DNOC (2-methyl-4-nitrophenol, 2-methyl-6-nitrophe-

nol and 2-methylphenol). However, none of these de-

gradation products were found above the impurity level

(<0.5 lg/l) from the mother product.

The BAM and bentazone results were not surprising

since such degradations have not been reported in the

literature. The recalcitrant behaviour of isoproturon was

contrary to the ®ndings of Johnson et al. (1998) who

found that isoproturon can be degraded under aerobic

conditions in a chalk aquifer.

Modelling was used to estimate the actual lag peri-

ods and degradation rates of the degraded pesticides

because the transport and the sorption parameters then

can be excluded. In the model, a lag period is de®ned as

the time from the beginning of the experiment to the

time where the concentration in the outlet starts to

decline. In the present study the residence time in the

columns was around 20 days. Consequently there is a

di�erence in the period of contact between sediment

and pesticides when looking at sediment near the inlet

or the outlet of the columns. The ``true'' environmental

lag period is therefore somewhere between the model

lag period and the model lag period minus the residence

time.

The DNOC breakthrough curve was modelled using

both a 0.-order and a 1.-order degradation model (Fig.

4). The 0.-order model simulated the observations better

than the 1.-order model, for all of the degraded pesti-

cides. After a model lag period of 80 days the degra-

dation of DNOC started with a 0.-order degradation

rate of 2.1 lg/l/day. Aerobic degradation of DNOC has

not previously been reported for in aquifers, but other

nitrophenols with similar molecular structures such as o-

and p-nitrophenol have been observed to be degraded

under aerobic conditions (Nielsen et al., 1996).

The model lag periods for the phenoxy acids ranged

between 16 and 33 days and 0.-order degradation rates

of 1.3±2.6 lg/l/day were found (Table 4). The literature

also suggests that degradation of phenoxy acids is likely

to occur after lag periods of signi®cant but varying

lengths and with rates on the same order of magnitude

(Agertved et al., 1992; Heron and Christensen, 1992;

Klint et al., 1993; Kuhlmann et al., 1995).

The degradation rates of DNOC and the phenoxy

acids were estimated assuming that degradation oc-

curred over the entire length of the columns. This was,

however, not the case throughout the experiment. After

an experimental period of 116 days (long after any of the

degradable pesticides has been detected in the outlet)

water samples were taken every 10 cm in the columns

which received 2,4-DD and MCPP. Analysis of these

samples showed that the pesticides were completely de-

graded within the ®rst 5 cm of the columns. Probably

degradation occurred over the entire length of the col-

umns at the beginning of the experiment and over time

the pesticide fronts in the columns receded. This indi-

cates that the estimated degradation rates more likely

represent the minimum degradation rates in these

columns.

The estimated degradation rates might not be the

true values for the environment due to the complexity of

microbial degradation and the lack of detailed infor-

mation of e.g. types and distribution of microorganisms.

However, the rate-estimates obtained are useful char-

acteristics of the environment especially when compared

to other environments such as top soil, where the rates

are much higher. Furthermore these rates can be used to

rank the degradability of the pesticides under relevant

conditions.

The HPLC method had a detection limit of 2±3 lg/l

and it was therefore not possible to state whether de-

gradation proceeded until concentrations were below the

EU drinking water standards of 0.1 lg/l. Assuming that

the 0.-order degradation model is valid across the entire

concentration range (25±0.1 lg/l) degradation in the

present experiment would likely have continued until

pesticide concentrations were below the EU standards.

However, several investigations have shown that de-

gradation kinetics are dependent on the substrate con-

centration level (Simkins and Alexander, 1984; Schmidt

et al., 1985).

While degradation rates for the degraded pesticides

were very similar, there was a variance in the lag periods.

The experiments were conducted using oxygen-saturated

1492 N. Tuxen et al. / Chemosphere 41 (2000) 1485±1494

groundwater (11.5 mg/l) and compared to the natural

oxygen content of 3±8 mg/l in the aquifer this may have

enhanced the degradation rates. The lag periods in col-

umns with sediment I were generally shorter compared

to columns with sediment II, indicating that sediment II

had a higher initial degradation potential. No relation-

ship was observed regarding the two sediments and de-

gradation rates.

MCPP was present in four columns: two columns

with pesticide solution A containing only 2,4-DD and

MCPP, and two columns with pesticide solution B

containing six pesticides. The results from the two pairs

of columns were very similar indicating that the presence

of other pesticides does not a�ect the MCPP degrada-

tion rate in a low concentration level. This suggests that

it may be possible to investigate several pesticides at one

time.

4. Conclusions

Column experiments were successfully conducted

over a period of 140 days to investigate transport,

sorption and degradation of seven pesticides in aerobic

aquifer material. The pesticides can be divided into four

groups with respect to their sorption capability: (1)

MCPP, dichlorprop, 2,4-DD and bentazone are all weak

acids which sorbed to a limited degree, or not at all with

Kd-values from 0±0.04 l/kg (0.11 l/kg for 2,4-DD in one

column); (2) DNOC, also a weak acid, showed a speci®c

and relatively strong sorption, probably to the clay

minerals kaolinite and illite with a Kd-value of 0.32 l/kg;

(3) isoproturon a non-polar compound that sorbed to

both the organic carbon content of the sediment and the

mineral surfaces (Kd-values ranging from 0.06 to 0.17 l/

kg), and (4) BAM a more polar compound, sorbed very

little or not at all. The sorption of isoproturon and

DNOC was successfully modelled with a one-site kinetic

sorption model.

Neither bentazone, isoproturon nor BAM were de-

graded during the experimental period. In contrast, all

the phenoxy acids and DNOC were degraded after lag

periods of 16±33 days for the phenoxy acids and 80 days

for DNOC. After the lag periods, degradations pro-

ceeded rapidly with 0.-order degradation rates of 1.3±2.6

lg/l/day. The lag periods were shortest in sediment I, but

there were no di�erences in the rates observed in the two

sediments. The presence of other pesticides in the solu-

tions had no e�ect on the degradation behaviour of

MCPP. None of the most probable degradation prod-

ucts were found at the column outlet.

The application of a reactive solute transport model

to a ¯ow-through system provided an opportunity to

separate the processes involved and quantify relevant

parameters such as pore water velocity, dispersivity, Kd-

values and degradation rate constants.

Acknowledgements

This study was funded by The Danish Environmental

Research Programme and the Technical University of

Denmark and is part of a research programme focusing

on pesticides in groundwater. Bent Skov, Jens Schaarup

Sùrensen, Mona Refstrup and Anja Foverskov all con-

tributed to the technical aspects of this project and their

work is gratefully acknowledged.

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