Full text of DOI:10.1002/etc.5620211213

Environmental Toxicology and Chemistry, Vol. 21, No. 12, pp. 2617–2622, 2002
2002 SETAC
Printed in the USA
0730-7268/02 $9.00 .00
᭧
ϩ
INVESTIGATION OF AN ONSITE WASTEWATER TREATMENT SYSTEM IN SANDY
SOIL: SORPTION AND BIODEGRADATION OF LINEAR ALKYLBENZENE SULFONATE
J
ON
D
OI,† KAY H. MARKS,‡ ALVARO J. D
ECARVALHO,*§ DREW C. MCAVOY,ͦͦ ALLEN M. NIELSEN,#
L
OUIS RAVETZ,†† and MANUEL L. CANO††
K
†Aqua Survey, 499 Point Breeze Road, Flemington, New Jersey 08822, USA
‡Roy F. Weston, 254 Welsh Pool Road, Exton, Pennsylvania 19341, USA
§The Soap and Detergent Association, 1500 K Street NW, Washington, DC 20005
ͦͦThe Procter & Gamble Company, 6110 Center Hill Avenue, Cincinnati, Ohio 45224, USA
#Sasol North America, 12024 Vista Parke Drive, Austin, Texas 78726, USA
††Shell Chemical Company, 3333 Highway 6 South, Houston, Texas 77082, USA
(
Received 31 October 2001; Accepted 28 May 2002)
Abstract—The objective of this work was to determine the sorptive and biodegradable characteristics of linear alkylbenzene
sulfonate (LAS) in a soil below a Florida, USA, septic system drainfield. Three distinct soil samples were collected from the septic
system drainfield study site. These soils were used in laboratory sorption and biodegradation studies. Different concentrations of
LAS were added, in radiolabeled and unlabeled forms, to a series of test vessels that contained upgradient groundwater and the
soils collected from the study site. The sorption test was designed to determine the partitioning of LAS between groundwater and
soil in each sample. Results indicated that the sorption distribution coefficient (
Kd) decreased from 4.02 to 0.43 L/kg and that the
rate of ultimate biodegradation (first-order rate constant, 1) decreased from 2.17 to 0.08/d with increasing distance (0.7–1.2 m
k
vertically below ground surface [BGS] and 0 to 6.1 m horizontally) from the drainfield. The three soils showed 49.8 to 83.4%
LAS mineralization (percentage of theoretical CO2) over 45- or 59-d test periods. These results demonstrate that subsurface soils
in this system have the potential to sorb and biodegrade LAS.
Keywords—Sorption
Biodegradation
Linear alkylbenzene sulfonate
Surfactants
Septic systems
INTRODUCTION
ditions collected below a septic system drainfield in Jackson-
ville (FL, USA).
This study was part of a program that investigated the fate
and transport of household cleaning product surfactants in on-
site wastewater treatment systems (OWTS) [1,2]. Sorption and
biodegradability characteristics are two important parameters
for determining the fate of a material in the soil environment.
The primary objective of this work was to determine the po-
tential for LAS to sorb and biodegrade in soil beneath and
downgradient of a septic tank drainage field. A secondary
objective was to determine the sorption distribution coeffi-
cients and biodegradation rate constants of LAS for incorpo-
ration into a predictive model for OWTS.
Sorption of organic compounds to natural soil is important
to their fate, bioavailability, and remediation. The principal
determinant of the degree of sorption of organic compounds
to natural soils is the soil organic matter content [3–5]. The
partitioning of an organic compound between a liquid and solid
phase once equilibrium conditions are established is generally
Biodegradation is an important loss mechanism for deter-
mining the fate of an organic compound in the soil environ-
ment. The biodegradation process reduces the exposure con-
centration in the soil environment to higher-level organisms,
thus improving the overall environmental safety of a chemical.
One method of assessing the biodegradability of an organic
compound is to conduct a CO2 production test [8–10]. In this
study, a 14C-labeled LAS compound was used to evaluate its
rate and extent of mineralization as was done previously to
determine the ultimate biodegradability of LAS in soil and
aquifer water beneath a septic tank field in Ontario, Canada
[8]. The LAS sorption and biodegradation data obtained from
this study were used as inputs to a predictive model for an
on-site wastewater treatment system [2].
MATERIALS AND METHODS
The methodologies used in this study for both sorption and
biodegradation were based on methods provided by the U.S.
Environmental Protection Agency (EPA) Toxic Substance
Control Act (TSCA) [9,11] and Organization for Economic
Cooperation and Development (OECD) [10,12]. These studies
were performed in compliance with all pertinent U.S. EPA
Good Laboratory Practice regulations [13].
expressed as a distribution coefficient (Kd). The relative dis-
tribution of an organic compound between solid and liquid
phases depends on the physical and chemical properties of the
organic compound and the solid phase.
The sorption of LAS has been previously reported on sed-
iment [6] and in soil and aquifer material beneath a septic tank
tile field in Ontario, Canada [7]. The current study was de-
signed to measure the sorptive capacity and distribution of
LAS to soils with different characteristics and climatic con-
Test materials
Both unlabeled and uniformly ring-labeled [14C] LAS (do-
decylbenzene sulfonic acid, sodium salt) were obtained from
Sasol North America (Austin, TX, USA). The molecular
weight of the LAS used in this study was 348 g/mole. The
* To whom correspondence may be addressed
(adecarvalho@sdahq.org).
2617

2618
Environ. Toxicol. Chem. 21, 2002
J. Doi et al.
Upgradient water collection
The groundwater used in this study was collected upgra-
dient of the septic system drainfield in well BW-1 (Fig. 1).
The groundwater was collected on January 16, 1996, packed
in ice, and shipped by overnight delivery to the testing lab-
oratory. The groundwater was refrigerated at approximately
4ЊC until use within 48 h.
Analytical methods
An aliquot of each composited soil sample was analyzed
for the following parameters: pH, cation exchange capacity,
organic carbon, organic matter, water content, water holding
capacity, texture (sand, silt, clay), and bulk density according
to Standard Methods of Soil Analysis [14,15].
Liquid scintillation counting (LSC) was used to quantify
the amount of 14C activity in the study samples. Stock solu-
tions, solvent extracts, and aqueous and combustion samples
were combined with Ultima Gold௡ scintillation cocktail (Pack-
ard Instruments, Meriden, CT, USA) and analyzed using a
Packard Instruments model A2500 Tri Carb LSC analyzer. The
radiochemical purity was determined by radioactive thin-layer
chromatography using a Bioscan System 200 TLC Imaging
Scanner (Bioscan, Washington, DC, USA) [16].
Fig. 1. Diagram of septic system site showing the three soil and
upgradient water collection points. SC
ground well.
ϭ soil core; BW ϭ back-
Sorption determination
The test apparatus for the sorption study consisted of 30-
ml Teflon
௢ centrifuge tubes. All tubes, except the LAS con-
specific activity of the 14C-radiolabeled material was 77.0
mg, the radiochemical concentration was 93.3 Ci/ml, and the
92% based on radioactive thin-
␮Ci/
trols (no soil added), received 7 g of dry-weight-equivalent
composited soil and 7 ml of upgradient groundwater. The LAS
controls received 7 ml of upgradient groundwater. Radiola-
beled or unlabeled LAS was dosed to duplicate tubes for each
␮
radiochemical purity was
layer chromatography.
Ͼ
soil at concentrations of 40, 200, 700, and 2,000
LAS was added to the soil blank tubes.
␮g/L. No
Soil sample collection
The soils used in this study were collected from the Jack-
sonville study site on January 16, 1996. Soil core samples
were collected at three locations below and downgradient
of the drainfield using a stainless-steel soil recovery auger
with polycarbonate sample liners (see Fig. 1 for diagram of
site and sampling locations). The samples were packed in
dry ice and shipped overnight to the testing laboratory. Soil
core 1 (SC-1) was a composite sample of unsaturated soil
collected from locations along the infiltrative surface of
trench 1 approximately 71 to 91 cm BGS. Soil core 2 (SC-
2) was a composite sample from the same boreholes as SC-
1, but the cores were taken from the saturated zone below
SC-1 from 122 to 142 cm BGS. Soil core 3 (SC-3) was a
composite sample of the soil approximately 6.1 m down-
gradient (southwest) of trench 1. The SC-3 core was taken
from the same relative elevation as SC-2 (122–142 cm BGS)
and was in the saturated zone at the time of sampling. The
multiple samples from each designated site were compos-
ited, passed through a 2-mm screen to remove coarse ma-
The test vessels were equilibrated for 3 h on a shaker table
rotating at 180 rpm. The vessels were then centrifuged for 40
min at 5,600 g. Triplicate aliquots of each centrate were as-
sayed by LCS. A mass balance was determined on one tube
for each concentration from each soil site by extracting the
soil with methanol. The mass of test substance present on the
soil was assayed by LCS. Linear alkylbenzene sulfonate sorbed
to the tube walls was extracted with methanol and measured.
Aliquots of the soil pellets from three test vessels were com-
busted to confirm the effectiveness of the methanol extraction.
Biodegradation determination
The test apparatus consisted of 500-ml Erlenmeyer glass
flasks each containing 100 g of dry-weight-equivalent soil
from each site. The 14C-labeled LAS was dosed to triplicate
flasks for each site at concentrations of 50 and 2,000
␮g/kg.
The LAS concentrations chosen were representative of the
range of LAS concentrations estimated in the saturated soil of
the septic system drainfield as measured by methylene blue
active substance [17]. The water content of each treatment was
adjusted to approximately 30% with deionized water [18].
During the study, the flasks were aerated continuously with
a CO2-free air source. The flasks were weighed to determine
water loss and adjusted daily with deionized water to maintain
30% water content. The CO2 evolved from the flasks was
trapped in a 1.5-N potassium hydroxide solution. Samples from
the base traps were periodically collected on days 0.25, 1, 3,
5, 7, 10, 14, 18, 21, 28, 35, and 45, with additional samples
on days 52 and 59 for site 3 to account for lag time. These
samples were assayed for 14C activity by LSC. After 45 d, the
terials, and then refrigerated at approximately 4
within 48 h.
ЊC until use
The
Kd for LAS to the study site soils was determined using
the following equation:
x
΂ ΃
m
Kd
ϭ
(1)
Ce
where
soil (
solution (mg/L).
x
ϭ
the mass of LAS sorbed (mg),
m ϭ the mass of
k
g), and Ce the equilibrium concentration of LAS in
ϭ

Sorption and biodegradation of LAS in sandy soil
Environ. Toxicol. Chem. 21, 2002
2619
Table 1. Soil characterization data
Study soils
Other typical soilsa
Loamy
sand
Parameter
SC-1b
SC-2
SC-3
Alfisol
Spodosol
4.0–5.0
Entisol
6.6–8.0
Ͼ10
1.0–4.0
—
—
—
—
—
pH
6.1
6.3
0.75
0.03
0.05
1.34
0.82
93.2
4.0
6.9
0.78
5.5–6.5
10–15
1.0–1.5
—
—
—
—
—
10–20
—
—
4.6
32.7
9.25
—
—
—
81.6
12.6
6.0
—
CECc (mequiv/100 g)
Organic carbon (%)
Organic matter (%)
WHCd (%) 1/3 bar
WHCd (%) 15 bar
Sand (%)
0.99
0.13
0.22
2.12
1.03
93.2
4.0
2.8
Ͻ
10
Ͻ
Ͻ
0.01
0.01
1.22
0.37
1.5–3.5
—
—
—
—
93.2
4.0
2.8
Sand
1.47
Silt (%)
Clay (%)
—
2.8
Sand
1.51
Յ
10
11–25
—
—
Soil classification
Sand
1.45
—
—
Bulk density (g/cm3)
—
a Soils used in U.S. Environmental Protection Agency and Organization for Economic Cooperation and
Development biodegradation studies.
b SC
ϭ
ϭ
soil core.
cation exchange capacity.
ϭ water holding capacity.
c CEC
d WHC
contents of the flasks containing soil from sites 1 and 2 were
acidified with 1 N hydrochloric acid (HCl) and aerated for 3
d. A final set of base trap samples was collected and analyzed
by LSC on day 48. The flasks from site 3 were acidified with
1 N HCL on day 59 and aerated for 3 d, and the base traps
were analyzed by LSC.
A mass balance was determined on one flask per each trip-
licate set. Each flask was centrifuged for 30 min at 5,600 g.
Triplicate aliquots (1 ml) of the centrate were assayed by LCS.
The weight of the remaining soil pellet was documented, and
five aliquots (0.1 g) were combusted using a sample oxidizer
(Packard Instruments model 307 oxidizer). Combustion of soil
pellets was performed to determine the amount of 14C radio-
activity in the soil [17].
(
K
d). A summary of the sorption data for the three soils col-
lected at the study site is shown in Table 2. The amount of
14C activity in the centrate increased or remained the same,
whereas the amount of 14C activity in the soil decreased from
site 1 to site 3. Consequently, the Kd values were lower in the
soil from site 3 when compared to the soil from site 1. Test
substance recoveries in these experiments ranged from 62 to
104%.
Biodegradation
The production of CO2 is an indication of the amount of
test substance that is ultimately biodegraded or mineralized by
microorganisms in a soil matrix test system. In this study, the
14C activity mass balance for all the soils tested averaged 103.5
To determine the extent of LAS ultimate biodegradation,
the 14CO2 was compared to the amount that could be theoret-
ically produced if all the carbon from LAS was mineralized
to 14CO2 and water (theoretical 14CO2 or T14CO2). To determine
the LAS mineralization rate, the 14CO2 data were analyzed with
a nonlinear first-order regression model of the form
Ϯ 6.0% and ranged between 92.2 and 109.5%. A summary of
the biodegradation data is presented in Table 3, while the 14CO2
production-versus-time curves are shown in Figures 2 and 3.
Examination of these biodegradation data results in four
observations. A direct correlation was observed between the
level of organic carbon content in the soils and the initial LAS
biodegradation activity. The organic carbon content decreased
Ϫ
k1(tϪc)
a
(1
Ϫ
e
)
for
for
t
t
Ͼ
c
c
from 0.13 to
Ͻ0.01% from site 1 to site 3 (Table 1), while the
Y
ϭ
(2)
Ά0
Յ
highest first-order biodegradation rates observed in these soils
where
a
ϭ
asymptote of curve (% T14CO2),
k
1 ϭ mineralization
lag time before
time (day), and cu-
Table 2. Soil sorption data
biodegradation rate constant (per day),
14CO2 production occurs (day),
mulative % T14CO2.
c ϭ
t
ϭ
Y ϭ
Site
no.
Dose
Centrate concn.
g/L)a
Solids concn. Kd (L/kg)
(
␮
g/L)
(
␮
(
␮
g/kg)
measured
RESULTS
SC-1b
SC-1
SC-1
SC-1
SC-1
SC-2
SC-2
SC-2
SC-2
SC-3
SC-3
SC-3
SC-3
40
40
200
700
2,000
40
200
700
2,000
40
12.67
11.58
15.62
13.31
130.8
452.5
1.23
1.15
2.79
3.53
4.02
0.92
0.87
1.09
0.89
0.81
0.85
0.83
0.43
The soils in this study were comprised primarily of sand with
smaller amounts of organic carbon and clay (Table 1). Higher
amounts of organic matter were found in soil collected closest
to the leaching field (SC-1 and SC-2). The organic carbon, clay
content, and cation exchange capacity values for the soils in
this study were markedly lower than the corresponding values
for the four typical soils used in U.S. EPA TSCA [11] and
OECD [13] soil biodegradation studies (Table 1).
46.86 (46.46)
128.3 (136.4)
345.4 (344.4)
21.62 (21.02)
86.36 (78.34)
275.9 (262.3)
836.0 (826.8)
15.74 (16.70)
75.76 (75.66)
273.7 (286.6)
1,387
19.98
75.02
301.4
746.0
12.82
64.06
228.3
476.8
200
700
2,000
Sorption
1,115
(1,057)
The amount of 14C activity measured in the centrate (liquid
phase after centrifugation) and in the soil methanol extracts
was used to determine the sorption distribution coefficients
a Values in parentheses are the aqueous concentration in
replicate test vessel.
␮g/L for a
b SC
ϭ soil core.

2620
Environ. Toxicol. Chem. 21, 2002
J. Doi et al.
Table 3. Summary of biodegradation data
Kinetic data (
Ϯ
standard error)a
Soil
site no.
Concn. LAS
Final %
T14CO2
Asymptote
% T14CO2
Rateb
(per day)
(
␮
g/kg)
Lag (d)
SC-1c
50
68.8
71.7
70.3
73.1
75.1
68.3
49.8
74.6
72.8
71.5
78.4
69.8
73.3
74.5
68.3
83.4
38.2
37.2
37.1
35.4
34.6
38.4
23.8
42.4
32.0
30.6
77.9
68.5
75.6
74.1
67.0
84.2
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
0.7
1.8
1.3
0.9
1.3
3.0
0.9
5.1
0.6
0.8
0.6
0.9
0.8
0.9
0.8
0.5
2.01
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
0.28
0.11
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
Ϯ
0.03
2.09
2.17
0.74
0.88
0.41
0.62
0.20
0.52
0.55
0.09
0.09
0.09
0.08
0.10
0.08
1.02
0.77
0.10
0.19
0.12
0.11
0.07
0.05
0.08
0.00
0.01
0.00
0.00
0.01
0.00
0.00
0.00
0.21
0.14
0.05
0.20
0.22
0.71
0.73
5.2
0.14
0.10
0.05
0.09
0.22
0.09
0.34
0.08
0.10
0.2
SC-1
SC-2
SC-2
2,000
50
2,000
SC-3
SC-3
50
5.1
6.6
9.1
9.1
0.3
0.2
0.2
0.2
2,000
9.8
0.1
a These parameters are calculated from the first order portions of the biodegradation curves.
b The data (site 1 at 2,000
␮g/kg and site 2 at 50 ␮g/kg) were not included in the respective analyses
because in both cases the final 14CO2 evolved was considerably less than other flasks of the same
treatment. No visual observations were made during test setup or throughout the sampling period to
explain the lower results. Knowledge of the site and interpretation of other biodegradation data de-
termined that these data points should be eliminated as outliers.
c SC
ϭ soil core.
decreased from 2.17 to 0.08/d, and the lag times increased
from 0.11 to 9.8 d (Table 3). An increase in LAS concentration
DISCUSSION
The ability of a soil to sorb and biodegrade organic com-
pounds is dependent on its characteristics. The soils in this
study were primarily sandy with organic matter, clay, or silt
content much less than typical soils within the United States
(Table 1). A direct correlation appears to exist with the amount
of organic content of a soil and its potential for sorption and
biodegradation as evidenced by lower sorption and biodeg-
radation activity at site 3 as compared to site 1, likely due to
an increase in the microbial population near the effluent source.
The sorption data from this study (Table 2) are consistent
with other observations reported in the literature. The sorption
from 50 to 2,000
␮g/kg negatively affected the initial LAS
biodegradation activity. In all soils, the initial first-order bio-
degradation rates stayed constant or decreased, and the lag
times increased at the higher LAS concentration (Table 3). The
LAS biodegradation patterns in soil from sites 1 and 2 were
biphasic, while the pattern in the soil from site 3 was mono-
phasic. Despite the initial differences in the LAS biodegra-
dation rates and lag times, the final amount of LAS mineral-
ization was extensive and very similar in all the soils with
final mineralization values (% T 14CO2) ranging from 49.8 to
83.4% with most (15 of 16)
Ͼ68% (Table 3).
Fig. 2. Biodegradation of linear alkylbenzene sulfonate (LAS) in a
Fig. 3. Biodegradation of linear alkylbenzene sulfonate (LAS) in a
sandy soil—14CO2 evolved (LAS concentrations: 50
␮
g/kg) for the
sandy soil—14CO2 evolved (LAS concentrations: 2,000
␮g/kg) for the
three sites. Each point is the average 14CO2 evolved from three flasks.
Insert shows the first-order kinetic portions of the biodegradation
curves along with the curve fits based on Equation 2.
three sites. Each point is the average 14CO2 evolved from three flasks.
Insert shows the first-order kinetic portions of the biodegradation
curves along with the curve fits based on Equation 2.

Sorption and biodegradation of LAS in sandy soil
Environ. Toxicol. Chem. 21, 2002
2621
of LAS to aquatic sediments was shown to more closely cor-
relate to the organic carbon content of the sediments than to
the LAS homolog tested, the pH or calcium concentration of
the test solution, or solids concentrations [6]. In soils beneath
a septic tank drainfield in Ontario, Canada, the sorption of
LAS was correlated with organic carbon and clay content in
the subsurface soils [7]. The Kd for LAS in that study ranged
from 1 to 20 L/kg for soils and 20 to 3,019 L/kg for sediments
[19,20].
The biodegradation data in this study are also consistent
with other reports in the literature. In a biodegradation study
of LAS in the previously mentioned septic tank drainfield in
Ontario, Canada, it was shown that the process of adaptation
of the resident microorganisms influenced the biodegradation
of LAS [8]. This same phenomenon was noted in this study
since, in each soil, longer lag periods were required for the
microbial populations to adapt to the higher LAS concentration
(Table 3).
The rapid mineralization of LAS in soils has been widely
reported [19–25]. First-order rate constants for LAS miner-
alization in pristine, agricultural, and sludge-amended soils
ranged from 0.02 to 1.25/d. Linear alkylbenzene sulfonate
sorbed to various soil constituents and mixed with either a
natural woodlot soil or a sludge-amended soil exhibited first-
order rates of 0.19 to 1.25/d with a maximum mineralization
amount of 48.5 to 96.2% T14CO2 [22,23], respectively. In an-
other study by Larson et al. [24], the first-order rate constants
were quite similar and averaged 0.03/d for the surface soils
and 0.04/d for the subsurface soils. The results of this exper-
iment indicate that soils immediately below the drainfield,
where biological activity would be high, had rates slightly
higher (2.10/d) than sludge-amended soils. Downgradient
soils, which do not receive appreciable LAS concentrations
and therefore would not have high biological activity, show
slower rates (0.09/d) comparable with previously reported val-
ues for pristine soils.
The data shown in Figures 2 and 3 indicate that two separate
kinetic events occurred in the soils from sites 1 and 2. Only
one kinetic event occurred with the soil from site 3. This
phenomenon has been described as biphasic mineralization.
The first phase is consistent with typical first-order biodeg-
radation, and the second phase is consistent with zero-order
biodegradation [26]. The zero-order phase represents the min-
eralization of carbon that is incorporated into biomass. This
phenomenon is known as biomass turnover or endogenous
respiration.
rate constants were highest in the unsaturated soil immediately
below the drainfield and decreased with increasing distance
from the drainfield. The amount of ultimate biodegradation
was quite high given the conditions of the soil, and the final
percentage of ultimate biodegradation remained relatively con-
stant for the three soil sites. The biodegradation pattern of sites
1 and 2 were consistent with biphasic mineralization, while
biodegradation in site 3 could be described by simple first-
order kinetics. These results demonstrate that both sorption
and biodegradation of LAS occur in the soils beneath the study
site and indicate that these parameters are important factors
in the removal of LAS in a subsurface soil system. Data from
this paper were subsequently used to validate a mathematical
model developed to predict the fate and transport of down-
the-drain household chemicals in septic systems [2,27].
Acknowledgement—The authors would like to thank John Hash and
Kathleen Crapo of the Weston Environmental Fate and Effect Lab-
oratory. We would also like to thank Ayres & Associates and A&L
Great Lakes Laboratory.
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