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WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 0
0043-1354/$ - see frodoi:10.1016/j.watres
�Corresponding auBedfordshire MK43
E-mail [email protected]@unilsean.o’connor@unil
journal homepage: www.elsevier.com/locate/watres
The behaviour of linear alkyl benzene sulphonate underdirect discharge conditions in Vientiane, Lao PDR
M.J. Whelana,�, R. Van Egmonda, I. Guymerb, J.O. Lacoursiered, L.M.B. Voughtd,C. Finnegana, K.K. Foxc, C. Sparhama, S. O’Connora, M. Vaughana, J.M. Pearsonb
aSafety and Environmental Assurance Centre, Unilever Colworth Laboratory, Sharnbrook, Bedfordshire MK44 1LQ, UKbSchool of Engineering, University of Warwick, UKcDepartment of Environmental Science, University of Lancaster, UKdDepartment of Mathematics and Science, University of Kristianstad, Sweden
a r t i c l e i n f o
Article history:
Received 13 April 2007
Received in revised form
19 June 2007
Accepted 24 June 2007
Available online 28 June 2007
Keywords:
Sewage
LAS
SPC
Tracing
Biodegradation
Ammonia
Lao PDR
nt matter & 2007 Elsevie.2007.06.059
thor. Department of Natu0AL, UK. Tel.: +44 1234 75: [email protected] (I. Guymer), jean.laever.com (C. Finnegan), kever.com (S. O’Connor), m
a b s t r a c t
Direct discharge of untreated sewage to surface waters is a common practice in many parts
of the world. However, relatively little is known about the behaviour of synthetic organic
pollutants under these conditions. This paper describes a sampling campaign designed to
track changes in water quality in a surface water system in Vientiane (Lao PDR) receiving
significant quantities of untreated waste water. The study was based on following in-
channel transport using a fluorescent tracer injected as a pulse, with a focus on the anionic
surfactant linear alkylbenzene sulphonate (LAS) and ammonia. Water samples were
collected at a number of stations with sampling times estimated to coincide with solute
time-of-travel. The reduction in LAS concentration with flow-time could be approximated
by first-order kinetics with a half life of about 7 h. Free ammonia concentrations decreased
more slowly than LAS and remained above the level believed to be toxic for sensitive
aquatic species along the entire channel. Changes in the ratios of LAS alkyl chain
homologues to total LAS concentrations suggest a preferential removal of longer chain
lengths. The role of biodegradation in the removal of LAS was confirmed by the presence of
LAS metabolites (sulphophenylcarboxylates, SPCs) which increased systematically (as a
fraction of LAS remaining) with flow-time.
& 2007 Elsevier Ltd. All rights reserved.
1. Introduction
Information about the environmental fate of many ‘down-
the-drain’ chemicals (such as the ingredients used in home
and personal care products and in pharmaceuticals) is
increasing and many monitoring studies have been per-
formed. However, most of these studies have been conducted
r Ltd. All rights reserved.
ral Resources, School of2975.c.uk (M.J. Whelan), [email protected] ([email protected]@unilever
in systems with temperate climates, receiving treated efflu-
ents (e.g. McAvoy et al., 1993; Waters and Feijtel, 1995; Tabor
and Barber, 1996; Fox et al., 2000). Relatively little is known
about chemical fate under direct discharge conditions (i.e.
where waste water is discharged to water courses without
treatment). Direct discharge is commonplace in many devel-
oping countries, particularly in the tropics (e.g. Eichhorn
Applied Sciences, Cranfield University, College Road, Cranfield,
[email protected] (R. Van Egmond),. Lacoursiere), [email protected] (L.M.B. Vought),m (K.K. Fox), [email protected] (C. Sparham),
.com (M. Vaughan), [email protected] (J.M. Pearson).
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WA T E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 0 4731
et al., 2001, 2002) and is typically associated with high levels
of suspended solids (SS), biochemical oxygen demand (BOD)
and free ammonia in receiving waters, in addition to
commercial chemicals. High levels of BOD can result in
significant depletion of dissolved oxygen (DO) concentrations,
which can be exacerbated by strong diurnal DO variations
resulting from algal and macrophytic photosynthesis and
respiration (e.g. Tadesse et al., 2004). This can produce major
ecological impacts in the immediate vicinity of emission
points, which often diminish with distance as pollutants are
removed by degradation, sedimentation and volatilisation
and as the rate of re-aeration exceeds the rate of DO removal
due to BOD oxidation (self-purification). There will commonly
be several waste water emission points along urban water
courses which may delay any amelioration of water quality
impacts until well beyond the urban fringe.
As far as ‘down-the-drain’ chemicals are concerned, there
are two main issues for direct discharge conditions. First,
conventional ecological risk assessments for specific chemi-
cals [which make a comparison between the predicted
environmental concentration (PEC) and the predicted no-
effect concentration (PNEC)] often have little meaning, since
the ecosystem in the receiving environment will already be
significantly impacted by other constituents present in raw
waste water. Secondly, the behaviour of the chemicals
themselves (e.g. their removal rates) may be affected by the
poor water quality (e.g. DO may be limiting to aerobic
degradation). An alternative and more appropriate risk
assessment model, based on the ‘‘impact zone’’ concept, has
been proposed for direct discharge conditions (Limlette III
Workshop, 1995; McAvoy et al., 2003). In this model, chemicals
are assessed in terms of their potential impact on river
recovery processes (e.g. microbial respiration and nitrifica-
tion) and in terms of their predicted concentration at the end
of an ‘‘impact zone’’ (IZ), within which the ecosystem is
impacted by pollutants such as free ammonia and BOD and
beyond which, it is not. It is, therefore, necessary to ascertain
the extent of chemical loss at the end of the IZ. McAvoy et al.
(2003) describe a monitoring and modelling study which was
conducted on a river receiving untreated sewage effluent in
the Philippines in which they applied IZ concepts to ascertain
the risks posed by the anionic surfactant, linear alkylbenzene
sulphonate (LAS). They showed that LAS was removed rapidly
from the water course and posed no risk to the ecosystem at
the end of the IZ (as defined by the concentration of free
ammonia). However, the Philippines study was conducted on
a relatively shallow river which was well oxygenated and
which may not be representative of all tropical direct
discharge situations.
This paper describes a monitoring study designed to track
the behaviour of LAS in a surface water system in Vientiane
(Lao PDR, Southeast Asia) which receives significant quan-
tities of untreated waste water. LAS was selected as a test
chemical for this study because it is a commonly used
ingredient in household detergents and has been studied
extensively in a number of different systems (e.g. Tabor and
Barber, 1996; Whelan et al., 1999; Gandolfi et al., 2000; Fox
et al., 2000; Eichhorn et al., 2001, 2002; McAvoy et al., 2003). It
is readily biodegradable and has been observed to degrade
rapidly under aerobic conditions in activated-sludge-type
sewage treatment (usually well in excess of 95% of the load,
Matthijs et al., 1997; Facchi et al., 2007) and in trickling filter-
type sewage treatment (removal rates of between 70% and
99% have been reported for UK plants by Holt et al., 1998). It
has also been shown to be effectively removed in both sewer
systems and rivers. In-sewer LAS removal rates reported in
the literature lie between 0% and 60% (Matthijs et al., 1995,
1997; Moreno et al., 1990; Holt et al., 1998; Boeije, 1999). The
range of in-stream half lives for LAS reported in a review by
Boeije (1999) is 0.40–116 h, depending on flow depth and
environmental conditions (e.g. temperature and DO concen-
tration). Typical half lives are of the order of 2–15 h for most
temperate rivers, depending on temperature, depth and the
condition of the channel bed and banks (Boeije et al., 2000).
The main objective of the study was to improve under-
standing of LAS behaviour relative to other water quality
indicators, particularly ammonia and DO. Specifically, the
following hypothesis was tested: LAS concentrations will not
fall below the LAS PNEC at the end of the IZ (defined as the
point at which free ammonia concentrations fall below the
PNEC for free ammonia). The study was conducted in a semi-
natural river channel which, after receiving the majority of
Vientiane’s waste water, flows through a rural area with very
low population density. In-channel solute time-of-travel was
followed using a fluorescent dye tracer (Rhodamine WT)
injected as a pulse. Discharge was also determined using the
velocity–area method. In addition to the analysis of LAS, the
presence of the principal LAS metabolites, sulphophenylcar-
boxylates (SPCs), was determined to confirm the role of
biodegradation in any changes in LAS concentration. SPCs are
formed by the biodegradation of LAS, mainly by o-oxidation
of the alkyl chain followed by b-oxidation which shortens the
chain lengths by two carbon atoms, respectively, to give a
wide range of homologues and isomers (Swisher, 1987;
Gonzalez-Mazo et al., 1997; Schleheck et al., 2004). Note that
SPCs are, themselves, biodegradable (e.g. Leon, et al., 2004).
They are less hydrophobic than LAS and are consequently
less toxic (e.g. the toxicity of C11 SPC was reported to be about
one order of magnitude lower than that of C11 LAS by Volpi
Ghirardini et al., 2001).
2. Methods
2.1. Site description
The study was conducted in the Houay Mak Hiao River, which
drains the urban area of Vientiane Capital City (formally
Vientiane Municipality) through the large That Luang wet-
land, Lao PDR. Location and key features are shown in Fig. 1.
The core of Vientiane (17157.70N, 102137.20E) is situated
between the left bank of the Mekong River and the That
Luang wetland area to the east. It has a population of
approximately 331,000 and a population density of about
134 people/km2 (Keosithamma, 2004). Detergent use in Vien-
tiane is believed to be similar to other parts of Asia. LAS is the
principal anionic surfactant used in most laundry detergents
and LAS concentrations measured in a pilot study conducted
by the authors (unpublished) in urban sewers typically range
between 2 and 5 mg L�1, which is similar to raw waste water
ARTICLE IN PRESS
0
LAO PDR
Gulf ofTonkin
LAO PDR
Gulf ofTonkin
A
B
C
DE
F
A
B
C
DE
F
b
LAO PDR
Gulf ofTonkin
LAO PDR
A
B
C
DE
F
A
B
C
DE
F
MAYANMAR
CHINA
VIETNAM
CHINA
Gulf of
Tonkin
VIENTIANE
THAILAND
Mekong
100 km50
a
Channel
Hong Xeng
That LuangWetland
1 km
Fig. 1 – (a) The regional location of the Lao PDR and Vientiane. (b) Detailed map showing the monitored channel and the
sampling locations.
Table 1 – Details of sampling stations used in the study including local name, distance along the channel from injection,number of samples (n) collected during the dye trace and the approximate sampling interval after the arrival of peakfluorescence
Stationcode
Station name anddescription
Distance from injection(km)
Number of samples and sampling interval afterdye peak
A That Luang Market Bridge 0 n ¼ 3, NA
B Bank upstream of Hong Xeng
channel
1.2 n ¼ 5, 30 min
C Bridge at Nong Nieng 2.63 n ¼ 5, 30 min
D Bank at Hongsouphap 5.09 n ¼ 7, 60 min
E Bridge at Ban Sok 7.2 n ¼ 8, 60 min
F Bridge at Na Kuai 9.74 n ¼ 4, 60 min
In the case of Station A, samples were collected 50 min before and 40 min after injection.
WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 04732
influent concentrations observed in Europe (e.g. Waters and
Feijtel, 1995; Holt et al., 1998). Mean annual rainfall is
1714 mm (www.bbc.co.uk/weather) which falls mainly in a
pronounced wet season (May to October). The average daily
minimum and maximum air temperatures are 20.5 and
30.8 1C, respectively (www.bbc.co.uk/weather). Most domestic
and commercial waste water is discharged (mainly via
rudimentary septic tanks) into open channels, all of which
flow eastwards, away from the Mekong towards the That
Luang wetland. The wetland, in turn, drains through the
Houay Mak Hiao river (approximately 65 km long) which
eventually leads to the Mekong.
The first 10 km of the Houay Mak Hiao river (from the That
Luang Market Bridge to Na Kuai) was selected as the study
section. Sampling was conducted at six stations between
20 and 30 March 2006. Details of each station are shown in
Table 1. There are two main waste water inflows into the river.
The first is via the That Luang wetland itself (from the Hong
Ke channel, draining via Station A) and the second is via the
Hong Xeng channel which joins the wetland channel just
downstream of Station B (see Fig. 1b). However, when
monitoring was conducted, the flow in the Hong Xeng
channel was negligible due to extensive damming and
construction works in that part of the drainage network.
There were, therefore, no significant known emissions of
waste water downstream of Station A and the system can be
regarded as having one single-point source. The discharge (by
the velocity–area method) at Station A at the time of injection
(07 h 52 min on the 26 March 2006) was 0.78 m3 s�1. Although a
significant fraction of this flow will be waste water, it will also
be augmented by urban storm runoff and groundwater
contributions to the That Luang wetland.
2.2. Sampling and dye tracing
In each experiment Rhodamine WT (Hubbard et al., 1982),
was injected into the flow as a pulse and its concentration
was monitored using a combination of continuously logged
fluorimeters (self-contained underwater fluorescence appa-
ratus: SCUFA) fixed in situ at several locations downstream
and spot measurements of fluorescence in discrete samples
using hand-held Picofluor fluorometers (Turner Designs Inc.).
The timing of sample collection at downstream stations was
based on the arrival of the dye peak—ascertained using the
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WA T E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 0 4733
hand-held fluorimeters and subsequently confirmed using
the SCUFA data. Samples were collected at the peak and at set
intervals after the peak in order to attempt to capture the
tracer centre of mass (centroid). The number of samples
collected at each station and the sampling interval used are
shown in Table 1. By staggering water sampling to coincide
with the arrival of the centroid, water quality changes during
solute transport can be quantified and chemical-loss rate
constants can be estimated (see Fox et al., 2000; McAvoy et al.,
2003).
All samples were collected by hand using a stainless-steel
container. Samples for LAS analysis were added to graduated
1 L glass bottles, which had been pre-washed with methanol.
Each bottle contained 30 mL of formalin (42% formaldehyde
solution) which was added as a preservative (e.g. Eichhorn et
al., 2002). The bottles were then transported to a local
laboratory in Vientiane, at most 3 h after sampling, where
they were extracted onto solid phase cartridges within 24 h.
Samples for other water quality variables were collected in
500 mL polyethylthene bottles and analysed on the same day
as collection. The temperature, conductivity and DO concen-
tration of the water at each sampling station were measured
in situ using hand-held probes (YSI 550A, YSI Hydrodata Ltd.,
Letchworth, UK).
2.3. Analysis of SPC and LAS
All samples were extracted locally in Vientiane using a
validated solid phase extraction (SPE) procedure (Unilever,
unpublished). No correction was made for the volume of
formalin added to the samples which will give a consistent
error for all samples. Validation of the preservation procedure
was conducted for LAS only and included comparison of
locally extracted concentrations with water sub-samples
preserved with 3% v/v formalin in Vientiane, transported via
courier and extracted in the UK. The correlation (data not
shown) was excellent (r2¼ 0.99), the slope of the best-fit line
was 1.06 and the average difference was less than 5% (n ¼ 37).
Samples were adjusted to pH 3 to facilitate the simultaneous
analysis of both SPC and LAS (Gonzalez-Mazo et al., 1997;
Eichhorn et al., 2002) before loading onto methanol-condi-
tioned C18 1 g/6 mL isolute solid phase cartridges (Kinesis,
Bolnhurst, UK). Cartridges were refrigerated in the dark until
transportation (under ambient conditions) back to a GLP-
compliant laboratory in the UK. SPE cartridges were subse-
quently eluted with methanol and analysed for four LAS
homologues (C10, C11, C12 and C13) using single-ion monitoring
(SIM) liquid chromatography/mass spectrometry (LC–MS).
Sample extracts were analysed using an Agilent 1100 LC-MS
with a 5 mm Varian Pursuit C18 150�2.0 mm column. A volume
of 10mL was injected onto the system with a flow rate of
0.25 mL min�1 and a column temperature of 30 1C. The mobile
phases were 5 mM ammonium formate (pH 3.0) (Mobile Phase
A) and acetonitrile (Mobile Phase B). The mobile phase
programme started with 25% Mobile Phase B rising to 100%
B over 25 min. This was held for 10 min and then equilibrated
back to the original conditions for 20 min. Analysis was
carried out using negative ion mode electrospray ionisation
(ESI) with a drying gas temperature of 300 1C, flowing at
7.0 L min�1, a nebuliser pressure of 35 psi and a capillary
voltage of 4000 V. The [M]� ions were monitored for each
analyte (m/z 297, 311, 325 and 339 for LAS C10 to LAS C13,
respectively, and 271, 285, 299, 313, 327, 341, 355 and 369 for
SPC C6 to SPC C13, respectively). Calibration curves for each
LAS homologue were constructed based on the known
distribution of the LAS standard for C10, C11, C12 and C13 i.e.
13.2%, 32.8%, 31.3%, 22.8% w/w, respectively. The limit of
detection (LOD) for total LAS was approximately 5mg L�1.
In the absence of an SPC standard, individual chromato-
gram peaks were identified by their molecular ion and the
peak area for each homologue was quantified against the
calibration curve of LAS C10. This procedure was taken from
Gonzalez-Mazo et al. (1997). Qualitative analysis of SPCs have
also been reported by Tabor and Barber (1996) and by Eichhorn
et al. (2001). As an example, a reconstituted ion chromato-
gram (RIC) of LAS is shown in Fig. 2, overlain with an RIC for
SPCs to show the relative intensities of the compounds in a
typical sample. Typical SIM chromatograms for SPCs are also
shown in Fig. 2. As the ions for SPCs C8–C11 and LAS C10–C13
are only 2 amu apart, the natural abundance 13C LAS isotope
[M+2]� also appears in Fig. 2 (as denoted). In the absence of
standards, the elution order of SPC and LAS species was
found to be consistent with those reported by other workers.
A peak area of 50 000 was set as the detection limit for SPC
homologues. When this was quantified against the LAS C10
calibration curve, it equated to a LOD of approximately
0.8mg L�1. No attempt was made (for either LAS or SPC) to
quantify the concentrations of different positional isomers in
each homologue. In any case, quantitative resolution of
structural fractions for SPCs would not have been possible
using the method employed in the absence of specific SPC
standards.
Quality control samples (duplicate samples and blanks
spiked with a 1000mg mL�1 LAS-in-methanol spiking solution
to concentrations in the range 200–1000mg L�1 in the local
laboratory in Vientiane) demonstrated total LAS recovery
ranged approximately from 54.9% to 78.8% (n ¼ 6) with good
recovery for all homologues. Although the LAS recovery was
lower for acidified compared with non-acidified samples,
these results are still considered to give quantitative LAS
concentrations in the samples analysed and the benefit of
having SPC extraction, in combination with LAS, outweighed
the slight reduction in recoveries. It should be noted that no
correction for the recovery of LAS was applied to the sample
results presented since spike recoveries were not performed
for every sample and because the recoveries were variable.
The error introduced by imperfect recovery is assumed to be
small and randomly distributed between samples. To com-
bine the extraction procedure for SPC and LAS, methanol was
not added to samples in the loading process. This is reflected
in a slightly lower recovery of the more hydrophobic species,
i.e. C13 (49.7–63.4%) compared to C10 (58.8–90.0%).
To facilitate sample processing in Lao PDR, all SPE equip-
ment (cartridges, vacuum manifold, vacuum pump and
tubing), methanol-washed bottles, LAS spiking solution and
hydrochloric acid were sent from the UK to a local agent.
Methanol (ACS, Merck), formalin (Minipiao Pharmacy) and
distilled water (Water Quality Monitoring Laboratory, Dept.
Irrigation, Ministry of Agriculture and Forestry, Lao PDR) were
obtained locally. The LAS concentration on distilled water
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0
*
26000
20000
15000
10000
5000
0
2 4 6 8 10 12 14 min
500000
400000
300000
200000
100000
0
M S
Res
pons
e
5 10 15 20 25Retention Time (mins)
LAS C10
LAS C11
LAS C12
LAS C13
M S
res
pons
e
10000
500
0
0
0
0
0
5000
1000
5
SPC C6
SPC C7 SPC C8 SPC C9 SPC C10SPC C11
SPC C12
SPC C13
10 15 20 25
5 10 15 20 25
5 10 15 20 25
5 10 15 20 25
5 10 15 20 25
5 10 15 20 25
5 10 15 20 25
5 10 15 20 25
Retention time (mins)
Fig. 2 – (a) Example RIC for a water sample from Station B showing LAS and SPC peaks. (b) Typical SIM chromatograms for
SPCs. Peaks marked with * are for natural abundance 13C LAS isotopes, [M+2]�.
WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 04734
blanks was always less than the LOD. LAS standards were
made up from commercial LAS paste (sodium salt) which was
fully characterised.
2.4. Analysis of nutrients, suspended solids and COD
In addition to LAS, water samples were analysed for nitrate
(NO3-N), ammonium (NH4-N) and chemical oxygen demand
(COD). SS, phosphate (PO4-P) and total phosphorus (TP) were
also determined on a limited number of samples from each
station. The NO2-N, NO3-N, NH4-N and COD analyses were
performed using colorimetric cuvette methods developed for
the Xion 500 Spectrophotometer (Hach-Lange GMBH, Dussel-
dorf, Germany). Although this system is intended for rapid
screening of these parameters, comparison of several sam-
ples analysed by the Water Quality Monitoring Laboratory
(MAF, Lao PDR) using standard methods showed excellent
agreement and suggested that the analyses were very
accurate.
SS were determined by filtering a known volume of water
through pre-burned, pre-weighed Whatman GFC filters. The
filters were dried at 60 1C for 24 h to get dry mass, weighed
and then re-weighed after burning for 2 h at 520 1C to get the
ash-free dry mass. PO4-P and TP concentrations were
determined in the Water Quality Monitoring Laboratory
(MAF, Lao PDR) in Vientiane. PO4-P was determined spectro-
photometrically (Murphy and Riley, 1962). TP concentrations
were determined by the same molybdenum blue method
after autoclave digestion with peroxodisulphate.
3. Results and discussion
The temporal pattern of fluorescence concentration observed
at each station is shown in Fig. 3. All data shown were
ARTICLE IN PRESS
WA T E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 0 4735
obtained from SCUFAs except for Station D where a technical
problem made the gathered data unreliable. The fluorescence
concentrations observed at Station F were low and there was
some uncertainty about the timing of the peak. The SCUFA
data suggested that the fluorescence peak occurred about 30 h
(1820 min) after injection but the hand-held data used to
guide sampling times suggested that the peak had occurred at
about 35 h. Since the timing of sampling was based on the
hand-held data, the latter value is used in the following
analysis, although it should be remembered that the total
time-of-travel to Station F may be overestimated.
Changes in the concentrations of total LAS and NH4-N with
time-of-travel during the dye trace are shown in Fig. 4. The
data show a very significant and clear decreasing trend in LAS
concentrations with travel time from Station A (from 513725
to 1472.9 mg L�1). The reduction in LAS concentration with
flow-time could be approximated very well by first-order
kinetics (r2¼ 0.99) with a rate constant of 0.096 h�1 and a half
life of about 7 h. This is slower than loss rate reported by
McAvoy et al. (2003) for a shallower river in the Philippines,
100
200
300
400
500
600
0
0
LA
S (
µg / L
)
500
Time to pea
1000
N
CLAS = 513e-0.0016t
R2 = 0.9922
Fig. 4 – Concentrations of mean total LAS (CLAS, solid line, left a
travel. t is time. Error bars show the mean7one standard devia
00 1500
B
C
DE F
B
C
DE F
Time After Injection (mins)
250020001000500
Flu
ore
scence C
oncentr
ation (
ppb)
300
400
200
100
Fig. 3 – Fluorescence concentration at different stations (B–F)
against time from injection (at Station A) during the main
dye tracing exercise. All data shown were obtained from
SCUFAs except for Station D which was obtained using a
hand-held fluorimeter.
but more rapid than losses reported for some temperate
rivers (e.g. Gandolfi et al., 2001). The rate of LAS degradation is
known to be affected by low DO concentrations (LAS is not
anerobically biodegradable) and this may have been a factor
in the system studied here. A strong diurnal pattern in DO
concentrations was observed (data not shown) along the
whole study reach, with day-time concentrations typically
about 50% of saturation (ca. 3 mg L�1) and night-time con-
centrations falling to o5% (ca. 0.3 mg L�1). However, this
variability is not reflected via any discernable variation in the
rate of LAS loss with time-of-travel. Several samples were
taken at each station (at set time intervals after the arrival of
the peak) to ensure that the centroid of the solute plume
(calculated a posteriori using the fluorescence data) was
sampled. However, there was very little difference in either
LAS or ammonium concentrations between samples collected
at any one station, except for ammonium at Station F. This
suggests that the system was in quasi-steady state (i.e. that
conditions were temporally constant in the short term) and
also lends confidence to the conclusion that significant water
quality changes were occurring with flow-time. It also
suggests that any errors associated with uncertainties about
the timing of the peak and centroid of fluoresence are
unlikely to affect the conclusions of the study.
River flow measurements indicate that discharge decreased
with distance downstream to site E (probably due to abstrac-
tions for irrigation) which suggests that water quality
changes were not due to dilution. The decrease in LAS
concentration which was observed was probably due to in-
stream degradation, although some sorption to sediment may
also have been responsible (Hand and Williams, 1987). While
it is likely that the LAS concentrations observed near the start
of the monitored section would have been high enough to
result in some ecological impact, the concentrations observed
beyond Station D (ca. 7 km from the main waste water inflow)
are lower than the PNEC proposed by Dyer et al. (2003) for LAS
(245mg L�1) derived from a species sensitivity distribution
approach. Note that although the LAS concentrations at
Stations A–C were higher than the aquatic PNEC, they were
2500
0
1
2
3
4
k (mins)
1500 2000
Am
monia
cal nitro
gen (
mg N
/ L
)3.5
2.5
1.5
0.5
H4 zero-order half life = 29.7 h
CNH4 = -0.001t + 3.5633
R2 = 0.8788
xis) and NH4-N (CNH4, dashed line, right axis) with time-of-
tion (see Table 1 for number of samples).
ARTICLE IN PRESS
0
50
100
150
200
250
A B C D E F
Concentr
ation (
ug/L
)
Station
C11
C10
C12
C13
Fig. 5 – Changes in the concentration of individual LAS homologues at different stations (see Table 1 for number of samples).
0A B C D E F
0
1
2
3
4
5
6
7
8
0
Fra
ctio
n o
f T
ota
l L
AS
0.6
0.5
0.4
0.3
0.2
0.1
Station
C13C12C11C10
Ra
tio
C10
: C
13
Time to Peak (mins)
500 1000 1500 2000 2500
Fig. 6 – (a) Changes in the fraction of different LAS
homologues at different stations and (b) Changes in the
ratio of C10:C13 homologues with time-of-travel (see Table 1
for number of samples).
WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 04736
well-below concentrations believed to inhibit respiration (e.g.
a NOEC for Pseudomonas putida of 35 mg L�1, normalised to the
C11.6 LAS structure, was reported by Feijtel et al., 1995) or
nitrification (unpublished internal Unilever data from an ISO
9509 [ISO, 1989] test suggest an EC0 450 mg L�1).
The distribution of LAS alkyl homologues in the samples
collected during the dye trace is shown in Fig. 5. At all
stations, this distribution is dominated by C11 and C12 LAS.
This reflects the distribution in the commercial material and
is consistent with observations reported by other workers
(e.g. Eichhorn et al., 2001).
In addition to a decrease in total LAS concentration with
distance downstream, significant and consistent changes in
the alkyl homologue distribution of total LAS were also
observed (Fig. 6). When expressed as a fraction of the total
LAS concentration, there was a decrease in longer chain LAS
homologues (C12 and C13) and an increase in shorter chain
homologues (C10 and C11) with time-of-travel. The average
ratio of C10:C13 for each station is also shown in Fig. 6. There is
a consistent increase in this ratio with time-of-travel which is
also consistent with observations reported in the literature of
a preferential removal of longer chain length material by both
biodegradation and adsorption (e.g. Hand and Williams, 1987;
Swisher, 1987; Matthijs and de Henau, 1987; Tabor and Barber,
1996; Eichhorn et al., 2001, 2002). Note that the C10:C13 ratio at
Station A (injection) was 1.14 (70.06) compared with a ratio
for the LAS standard of 0.6, suggesting that there was some
LAS degradation during transport from the main waste water
inflow to the That Luang wetland via the Hong Ke channel.
Unpublished data from previous studies conducted by the
authors in urban drainage channels in Vientiane suggest that
the rate of LAS biodegradation within the city environment
may be limited by low DO concentrations.
The concentration of both NH4-N and NO3-N increased with
time-of-travel to Station C and then decreased (Figs. 4 and 9).
Whilst this could have been due to an additional unknown
point-source input between Stations B and C, it is more likely
to have resulted from nitrogen transformations occurring
within the system, since there were no concurrent increases
in COD or LAS concentrations. Overall, the trend in NH4-N
concentration appeared to be approximately linear, suggest-
ing zero-order kinetics in the concentration range observed
(i.e. the rate of change in concentration is independent of
concentration) with a rate constant of about 0.06 mg N L�1 h�1
(r2¼ 0.88). Many authors have reported that nitrification can
be represented by Michaelis–Menten kinetics (e.g. Shieh and
La Motta, 1979) which would be consistent with zero-order
kinetics at high concentrations. However, the dynamics of
nitrogen in this system are likely to be complex with
ARTICLE IN PRESS
0A B C D E F
SP
C c
onc (
ug/L
) or
Ratio (
%)
250
200
150
100
50
av SPC/LAS %
SPC conc (ug/L)
Station
Fig. 7 – Change in total SPC concentration and in the ratio of
SPC: LAS concentration in dye trace samples. Error bars
show mean7one standard deviation (see Table 1 for
number of samples).
WA T E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 0 4737
ammonium concentration representing a balance between
formation (from the mineralisation of organic nitrogen) and
losses (due to volatilisation, plant and microbial uptake,
sorption to solids and nitrification). Indeed, mass balance
models based on assumptions of first-order kinetics for N
mineralisation and first-order kinetics for nitrification (e.g.
Chapra, 1997) can generate temporal changes in NH4-N
similar to that observed in this study. Nevertheless, there
are probably too many uncertainties to simulate the dy-
namics of nitrogen with confidence. Unionised (free) ammo-
nia represented approximately 0.89–0.98% of the total
ammonia concentration in this system (pH 6.89–7.17 average
7.04). Average free ammonia concentrations were estimated
to range between 44mg L�1 (at Station C ) and 15mg L�1 (at
Station F). For comparison, a toxicity threshold (PNEC) of
10mg L�1 for unionised ammonia was reported by Dyer et al.
(2003) based on a species sensitivity distribution approach,
while a value of 25mg L�1 has been reported by Alabaster and
Lloyd (1980) for the protection of freshwater fisheries based
on toxicity to salmonids. In the IZ model, the PEC for the
chemical of interest is calculated, and compared with the
PNEC, at the end of the IZ (i.e. the point in the system where
the toxicity of free ammonia falls below the ammonia PNEC).
The data reported here, suggest that the end of the IZ is
probably downstream of Station F (410 km from the main
waste water inflow), although its exact location is likely to
oscillate, depending on flow conditions. At Station F the
average concentration of total LAS was 14mg L�1 (72.9 mg L�1)
which is similar to levels observed in surface waters in
Western Europe receiving treated waste water (e.g. Schulze
et al., 1999; Schulze and Matthies, 2001; Holt et al., 2003) and
much lower than the PNEC for LAS proposed by Dyer et al.
(2003). The concentration at Station F was approximately 2.7%
of the concentration at the dye injection point (Station A).
When positioned within the IZ framework, this suggests that,
under the conditions studied, up to 97% of LAS may be
removed before the system is predicted to recover from high
ammonia levels. However, it should be noted that system-
specific ecological impacts (via bio-monitoring) were not
rigorously assessed in this study. It is important to recognise
that different ecosystems will be affected to different degrees
by individual stressors and that ecological impairment does
not necessarily result from concentrations above the PNEC.
Eichhorn et al. (2002) also observed a significant reduction
in LAS concentrations in a Brazilian river, receiving untreated
emissions of waste water, although they did not measure
time-of-travel or stagger sampling times accordingly. Mea-
surements of SPCs suggest that much of the observed loss
was due to in-stream biodegradation.
The estimated concentrations of total SPC in the samples
analysed are shown in Fig. 7 and the concentrations of each
SPC alkyl homologue, as quantified using the LAS C10
calibration curve, are shown in Fig. 8. Overall, there was a
slight decrease in total SPC concentration over the study
reach. However, there was a systematic increase in the ratio
of total SPC to total LAS concentration, particularly at the
lowest section of the studied reach (Station F). This pattern
confirms the inference from changes in total LAS and the LAS
C10: C13 ratio, that much of the observed LAS loss in this
system was due to biodegradation and is consistent with the
observations made by others (e.g. Eichhorn et al., 2002). The
concentration of SPCs represents a balance between forma-
tion (via LAS oxidation) and degradative loss (Leon et al.,
2004). The progressive increase in the SPC to LAS ratio with
time-of-travel implies that, in this system, the primary
degradation of LAS is more rapid than that of the SPCs.
Qualitatively, this confirms the laboratory-derived degrada-
tion kinetic data reported by Leon et al. (2004) for seawater
which suggested that the half life of total SPC was several
times greater than that for LAS.
In general, the distribution of SPC concentrations is
dominated by C7–C10 homologues (and, in particular, by C8
and C9 downstream of Station C: Fig. 8). This is consistent
with observations reported elsewhere and is probably due to
the fact that longer chain SPCs are converted relatively
rapidly to shorter chain lengths but with a reduced rate of
conversion with progressive chain shortening, resulting in a
high proportion of mid-length homologues (e.g. Eichhorn
et al., 2001).
When expressed as a fraction of the total SPC concentration
(data not shown), there appears to be a slight increase in
shorter chain SPCs (C6–C9) and a slight decrease in longer
chain homologues (C10–C13) with increasing flow-time. This
pattern reflects the relative change in the homologue
distribution observed for LAS and may be due to higher
losses by degradation and (possibly) sorption for the more
hydrophobic SPC species (although all SPCs are much less
hydrophobic than C10 LAS).
Changes in other water quality variables, expressed in
absolute terms and relative to LAS concentration at each
station in the wetland channel are shown in Fig. 9. The data
show a decrease in the concentrations of NH4-N and COD, an
increase in NO2-N concentration and a variable NO3-N
concentration with increasing flow-time. The variability in
concentrations between samples collected at different times
at the same station is generally much lower than between
station variations. The ratios of LAS to each water quality
determinant show a consistent decrease with distance from
injection. This suggests that the rate decrease in LAS
concentration is systematically higher than for ammonia
ARTICLE IN PRESS
0
0
0
0
5
0
2
4
6
8
0
1
2
3
4
5
6
0
0
NH
4 c
on
c (
mg
N /
L)
500
Time to Peak (mins)
1000 1500 2000 2500 0 500
Time to Peak (mins)
1000 1500 2000 2500
0 500
Time to Peak (mins)
1000 1500 2000 2500
NO
2 c
on
c (
mg
N /
L)
0.09
0.08
0.07
0.06
0.05
0.04
0.03
0.02
0.01
0.00
Time to Peak (mins)
500 1000 1500 2000 2500
NH4-N
NO2-N
LA
S :
NH
4-N
ra
tio
160
140
120
100
80
60
40
20
LA
S :
NO
2-N
ra
tio
25000
20000
15000
10000
5000
CO
D c
on
c (
mg
/ L
)
50
45
40
35
30
25
20
15
10
NO
3 c
on
c (
mg
N /
L)
1.8
1.6
1.4
1.2
1.0
0.8
0.6
0.4
0.2
0.0
LA
S :
NO
3-N
ra
tio
900
800
700
600
500
400
300
200
100
LA
S:
CO
Dra
tio
18
16
14
12
10
COD
NO3-N
Fig. 9 – Changes in the concentration of NH4, COD, NO2 and NO3 with time of travel (solid lines, left axis) and the ratio of the
LAS concentration to the concentration of each determinant (dashed lines, right axis). Error bars show mean71 SD (see Table
1 for number of samples).
0
2
4
6
8
C6
C7
C8
C9
C10
C11
C12
C13
A B C D E F
SP
C c
oncentr
ation (
µg/L
)
14
12
10
Station
Fig. 8 – Change in estimated average absolute concentrations of different SPC alkyl homologues in samples collected from
different stations.
WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 04738
and COD, which are important markers of waste water
contamination (note that nitrate was not determined on
samples collected from Station F). Data on SS showed no
significant trend with time-of-travel. SS concentrations
ranged from 19.9 mg L�1 (at Station B) to 46 mg L�1 (at Station
A) and were generally around 30 mg L�1. The organic matter
content of SS ranged between 34% and 59% and was typically
about 35%. There was an apparent reduction in both PO4-P
and TP concentrations with distance downstream from 0.49
(TP ¼ 0.59) mg L�1 (at Station A) to 0.24 (TP ¼ 0.39) mg L�1 (at
Station E) but the concentration of both determinants
increased to 0.47 (TP ¼ 0.57) mg L�1 at Station F.
4. Conclusions
In many parts of the world, domestic waste water is dis-
charged untreated into the environment. However, relatively
ARTICLE IN PRESS
WA T E R R E S E A R C H 4 1 ( 2 0 0 7 ) 4 7 3 0 – 4 7 4 0 4739
little is known about the fate and potential effects of
pollutants under these conditions. This paper describes a
sampling campaign designed to track the behaviour of LAS
and ammonia in a surface water system in Vientiane (Lao
PDR) receiving significant quantities of untreated sewage. The
study was based on following in-channel solute transport
using a tracer, injected as a pulse, and the collection of
samples downstream coincident with time-of-travel. The
data show that LAS is removed rapidly in the study channel
(from 513725 to 1472.9 mg L�1), confirming earlier work (e.g.
Fox et al., 2000; McAvoy et al., 2003) which suggests that LAS is
removed from receiving waters more rapidly than ammonia.
The change in LAS concentrations with time-of-travel could
be fitted using first-order kinetics with a half life of about 7 h.
When positioned within the IZ framework, the concentration
of LAS is likely to be much lower than the PNEC by the time
the system is predicted to recover from high concentrations
of free ammonia (i.e. NH3o10 mg L�1). Although it is difficult to
generalise, the results suggest that aquatic concentrations of
readily biodegradable substances, such as LAS, may be
reduced by as much as 97% within the IZ under tropical
direct discharge conditions. Qualitative analysis of benthic
invertebrates collected from this system by the authors
suggested that the in-stream ecology at Stations A, B and C
(ca. 4 km from the main waste water inflow) was heavily
impacted, with samples dominated by Chironomidae and
Oligochaeta. It also indicated that ecological quality improved
downstream, since the fauna was more diverse and a number
of species of snail were found at Stations D–F. Fish were also
abundant at these latter stations. However, in the absence of
more detailed biomonitoring data (and in the absence of
baseline reference site information), the relevance of the
PNECs assumed here (for both LAS and ammonia) as thresh-
olds for ecological impairment in this particular system is
uncertain. Moreover, LAS and ammonia are only two of
several factors which could affect the ecosystem, including
high SS concentration and a cyclic depression of DO at night
(which was most marked at Stations A–C). Further work is
required with different substances (including those which are
not as rapidly biodegradable as LAS) to ascertain the extent to
which these conclusions can be generalised. Ideally, any
future studies should also attempt to quantify the extent of
any ecological impairment in the IZ in order to ascertain the
validity of the findings for risk assessment.
Acknowledgements
The authors would like to thank Mrs. Phayvanh Bandavong
and her staff from the Water Quality Monitoring Laboratory
(Dept. of Irrigation, Ministry of Agriculture and Forestry) for
their expert assistance, Dr Andy Shaw for assistance in the
field and Mr. Lieng Khamsivilay, Director of the Living Aquatic
Research Centre (Ministry of Agriculture and Forestry) for
allowing the participation of Mrs. Khampheng Homsombath
and Mr. Onsy Ssuokelavong, our accomplished and dedicated
field associates. This project would not have been possible
without the facilitation and support of Mr. Sourasay Phou-
mavong, Deputy Director General of the Lao National Mekong
Committee Secretariat. Dr. Stuart Marshall from Unilever
provided valuable comments on the manuscript.
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