INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY
ENVIRONMENTAL HEALTH CRITERIA 169
LINEAR ALKYLBENZENE SULFONATES
AND RELATED COMPOUNDS
This report contains the collective views of an international group
of experts and does not necessarily represent the decisions or the
stated policy of the United Nations Environment Programme, the
International Labour Organisation, or the World Health Organization.
First draft prepared at the National Institute of Health Sciences,
Tokyo, Japan, and the Institute of Terrestrial Ecology, Monk's Wood,
United Kingdom
Published under the joint sponsorship of the United Nations
Environment Programme, the International Labour Organisation, and
the World Health Organization
World Health Organization
Geneva, 1996
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venture of the United Nations Environment Programme, the
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chemicals.
WHO Library Cataloguing in Publication Data
Linear Alkylbenzene Sulfonates and Related Compounds.
(Environmental health criteria ; 169)
1.Alkane sulfonates - adverse effects 2.Environmental exposure
3.Guidelines I.Series
ISBN 92 4 157169 1 (NLM Classification: QU 98)
ISSN 0250-863X
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CONTENTS
WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR LINEAR
ALKYLBENZENE SULFONATES AND RELATED COMPOUNDS
NOTE TO READERS OF THE CRITERIA MONOGRAPHS
ENVIRONMENTAL HEALTH CRITERIA FOR LINEAR ALKYLBENZENE SULFONATES AND
RELATED COMPOUNDS
1. OVERALL SUMMARY, EVALUATION, AND RECOMMENDATIONS
1.1. Identity and analytical methods
1.2. Sources of human and environmental exposure
1.3. Environmental concentrations
1.3.1. Linear alklylbenzene sulfonates
1.3.2. alpha-Olefin sulfonates and alkyl sulfates
1.4. Environmental transport, distribution, and transformation
1.4.1. Linear alklylbenzene sulfonates
1.4.2. alpha-Olefin sulfonates
1.4.3. Alkyl sulfates
1.5. Kinetics
1.6. Effects on experimental animals and in vitro
test systems
1.7. Effects on humans
1.8. Environmental effects
1.8.1. Linear alklylbenzene sulfonates
1.8.1.1 Aquatic environment
1.8.1.2 Terrestrial environment
1.8.1.3 Birds
1.8.2. alpha-Olefin sulfonates
1.8.2.1 Aquatic environment
1.8.2.2 Terrestrial environment
1.8.3. Alkyl sulfates
1.8.3.1 Aquatic environment
1.8.3.2 Terrestrial environment
1.9. Human health risk evaluation
1.10. Evaluation of effects on the environment
1.11. Recommendations for protection of human health
and the environment
1.12. Recommendations for further research
A. Linear alkylbenzene sulfonates and their salts.
A1. SUMMARY
A2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND
ANALYTICAL METHODS
A2.1 Identity (sodium salt)
A2.2 Physical and chemical properties
A2.3 Analysis
A2.3.1 Isolation
A2.3.2 Analytical methods
A2.3.2.1 Nonspecific methods
A2.3.2.2 Specific methods
A3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
A3.1 Natural occurrence
A3.2 Anthropogenic sources
A3.2.1 Production levels and processes
A3.2.2 Uses
A4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
Section summary
A4.1 Transport and distribution between media
A4.1.1 Wastewater treatment
A4.1.2 Surface waters, sediments, and soils
A4.1.3 Fate models
A4.2 Environmental transformation
A4.2.1 Biodegradation
A4.2.1.1 Aerobic degradation
A4.2.1.2 Anaerobic degradation
A4.2.2 Abiotic degradation
A4.2.2.1 Photodegradation
A4.2.2.2 Cobalt-60 irradiation
A4.2.3 Bioaccumulation and biomagnification
A4.2.3.1 Aquatic organisms
A4.2.3.2 Terrestrial plants
A5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
Section summary
A5.1 Environmental levels
A5.1.1 Wastewater, sewage effluent, and sludge
A5.1.2 Sediment
A5.1.3 Surface water
A5.1.4 Soil and groundwater
A5.1.5 Drinking-water
A5.1.6 Biota
A5.2 Environmental processes that influence concentrations
of linear alkylbenzene sulfonates
A5.2.1 Changes in chain length distribution during
environmental removal of linear alkylbenzene
sulfonates
A5.2.2 Specification of linear alkylbenzene sulfonates
in surface waters
A5.3 Estimation of human intake
A6. KINETICS
Section summary
A6.1 Absorption, distribution, and excretion
A6.2 Biotransformation
A7. EFFECTS ON LABORATORY MAMMALS AND IN VITRO TEST SYSTEMS
Section summary
A7.1 Single exposures
A7.2 Short-term exposure
A7.2.1 Mouse
A7.2.2 Rat
A7.2.2.1 Administration in the diet
A7.2.2.2 Administration by gavage
A7.2.2.3 Dermal application
A7.2.2.4 Subcutaneous injection
A7.2.3 Guinea-pig
A7.2.4 Monkey
A7.3 Long-term exposure; carcinogenicity
A7.3.1 Mouse
A7.3.1.1 Administration in the diet
A7.3.1.2 Administration in the drinking-water.
A7.3.2 Rat
A7.3.2.1 Administration in the diet
A7.3.2.2 Administration in the drinking-water.
A7.3.2.3 Administration by gavage
A7.3.2.4 Dermal application
A7.4 Skin and eye irritation; sensitization
A7.4.1 Studies of skin
A7.4.2 Studies of the eye
A7.5 Reproductive toxicity, embryotoxicity, and teratogenicity
A7.6 Mutagenicity and related end-points
A7.6.1 Studies in vitro
A7.6.2 Studies in vivo
A7.7 Special studies
A7.7.1 Studies in vitro
A7.7.2 Biochemical effects
A8. EFFECTS ON HUMANS
Section summary
A8.1 Exposure of the general population
A8.2 Clinical studies
A8.2.1 Skin irritation and sensitization
A8.2.2 Effects on the epidermis
A8.2.3 Hand eczema
A8.2.4 Occupational exposure
A8.2.5 Accidental or suicidal ingestion
A9. EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND THE FIELD
Section summary
A9.1 Effect of chain length on the toxicity of linear
alkylbenzene sulfonates
A9.2 Microorganisms
A9.3 Aquatic organisms
A9.3.1 Aquatic plants
A9.3.1.1 Freshwater algae and cyanobacteria
A9.3.1.2 Marine algae
A9.3.1.3 Macrophytes
A9.3.2 Aquatic invertebrates
A9.3.2.1 Acute toxicity
A9.3.2.2 Short-term and long-term toxicity
A9.3.2.3 Biochemical and physiological effects
A9.3.3 Fish
A9.3.3.1 Acute toxicity
A9.3.3.2 Chronic toxicity
A9.3.3.3 Biochemical and physiological effects
A9.3.3.4 Behavioural effects
A9.3.3.5 Interactive effects with other
chemicals
A9.3.4 Amphibia
A9.3.5 Studies of the mesocosm and communities
A9.3.6 Field studies
A9.3.7 Toxicity of biodegradation intermediates and
impurities of linear alkylbenzene sulfonates
A9.3.7.1 Individual compounds
A9.3.7.2 Effluents
A9.4 Terrestrial organisms
A9.4.1 Terrestrial plants
A9.4.2 Terrestrial invertebrates
A9.4.3 Birds
B. alpha-Olefin sulfonates
B1. SUMMARY
B2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND
ANALYTICAL METHODS
B2.1 Identity
B2.2 Physical and chemical properties
B2.3 Analytical methods
B3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
B3.1 Natural occurrence
B3.2 Anthropogenic sources
B3.2.1 Production levels and processes
B3.2.2 Uses
B4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
Section summary
B4.1 Transport and distribution between media
B4.2 Biotransformation
B4.2.1 Biodegradation
B4.2.1.1 Aerobic biodegradation
B4.2.1.2 Anaerobic degradation
B4.2.2 Abiotic degradation
B4.2.3 Bioaccumulation and biomagnification
B4.3 Interaction with other physical, chemical, and
biological factors
B5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
B6. KINETICS
Section summary
B6.1 Absorption, distribution, and excretion
B6.2 Biotransformation
B7. EFFECTS ON LABORATORY MAMMALS AND IN VITRO TEST SYSTEMS
Section summary
B7.1 Single exposures
B7.2 Short-term exposure
B7.3 Long-term exposure; carcinogenicity
B7.3.1 Mouse
B7.3.2 Rat
B7.4 Skin and eye irritation; sensitization
B7.5 Reproductive toxicity, embryotoxicity, and teratogenicity
B7.6 Mutagenicity and related end-points
B7.7 Special studies
B8. EFFECTS ON HUMANS
Section summary
B8.1 Exposure of the general population
B8.2 Clinical studies
B8.2.1 Skin irritation and sensitization
B8.2.2 Effect on the epidermis
B8.2.3 Hand eczema
B8.2.4 Accidental or suicidal ingestion
B9. EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND THE FIELD
Section summary
B9.1 Microorganisms
B9.2 Aquatic organisms
B9.2.1 Aquatic plants
B9.2.2 Aquatic invertebrates
B9.2.3 Fish
B9.3 Terrestrial organisms
B9.3.1 Terrestrial plants
B9.3.2 Terrestrial invertebrates
B9.3.3 Birds
C. Alkyl sulfates
C1. SUMMARY
C2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND
ANALYTICAL METHODS
C2.1 Identity
C2.2 Physical and chemical properties
C2.3 Analysis
C2.3.1 Isolation
C2.3.2 Analytical methods
C3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
Section summary
C3.1 Natural occurrence
C3.2 Anthropogenic sources
C3.2.1 Production levels and processes
C3.2.2 Uses
C4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
Section summary
C4.1 Transport and distribution between media
C4.2 Biotransformation
C4.2.1 Biodegradation
C4.2.1.1 Biodegradation pathway; mechanism
C4.2.1.2 Biodegradation in the environment
C4.2.1.3 Anaerobic degradation
C4.2.2 Abiotic degradation
C4.2.3 Bioaccumulation and biomagnification
C4.3 Interaction with other physical, chemical,
and biological factors
C4.4 Ultimate fate following use
C5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
Section summary
Environmental levels
C6. KINETICS
Section summary
C6.1 Absorption, distribution, and excretion
C6.2 Biotransformation
C7. EFFECTS ON LABORATORY MAMMALS AND IN VITRO TEST SYSTEMS
Section summary
C7.1 Single exposures
C7.2 Short-term exposure
C7.2.1 Rat
C7.2.1.1 Administration in the diet
C7.2.1.2 Administration in the drinking-water
C7.2.1.3 Dermal application
C7.2.2 Rabbit
C7.3 Long-term exposure; carcinogenicity
C7.3.1 Mouse
C7.3.2 Rat
C7.3.2.1 Administration in the diet
C7.3.2.2 Administration in the drinking-water
C7.4 Skin and eye irritation; sensitization
C7.4.1 Local irritation
C7.4.1.1 Skin
C7.4.1.2 Eye
C7.4.2 Skin sensitization
C7.5 Reproductive toxicity, embryotoxicity, and teratogenicity
C7.6 Mutagenicity and related end-points
C7.7 Special studies
C8. EFFECTS ON HUMANS
Section summary
C8.1 Exposure of the general population
C8.2 Clinical studies
C8.2.1 Skin irritation and sensitization
C8.2.2 Effects on the epidermis
C8.2.3 Hand eczema
C8.2.4 Accidental or suicidal ingestion
C9. EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND THE FIELD
Section summary
C9.1 Microorganisms
C9.2 Aquatic organisms
C9.2.1 Aquatic plants
C9.2.1.1 Freshwater algae
C9.2.1.2 Macrophytes
C9.2.2 Aquatic invertebrates
C9.2.3 Fish
C9.2.4 Tests in biocenoses
C9.3 Terrestrial organisms
APPENDIX I
REFERENCES
RESUME
RESUMEN
WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR LINEAR
ALKYLBENZENE SULFONATES AND RELATED COMPOUNDS
Members
Dr R.S. Chhabra, National Institutes of Health, Institute of
Environmental Health Sciences, Research Triangle Park, North
Carolina, USA
Dr A. Granmo, University of Göteborg, Marine Research Station at
Kristineberg, Fiskebackskil, Sweden
Ms K. Hughes, Priority Substances Section, Health and Welfare
Canada, Ottawa, Ontario, Canada
Mr H. Malcolm, Institute of Terrestrial Ecology, Huntingdon, United
Kingdom
Dr E. van der Plassche, Toxicology Advisory Centre, National
Institute of Public Health and Environmental Protection,
Bilthoven, Netherlands
Dr J. Sekizawa, Division of Information on Chemical Safety, National
Institute of Hygienic Sciences, Tokyo, Japan
Ms R. Takei, Research Planning and Administration Department, Lion
Corporation, Tokyo, Japan
Dr D.G. Van Ormer, Health Effects Division, Office of Pesticides
Programs, Environmental Protection Agency, Washington DC, USA
Professor P.N. Viswanathan, Industrial Toxicology Research Centre,
Lucknow, India
Representatives/Observers
IUTOX
Dr P. Montuschi, Department of Pharmacology, Catholic University of
the Sacred Heart, Rome, Italy
CEFIC
Dr J.L. Berna, Petresa, Madrid, Spain (20-21 October)
Dr L. Cavalli, Enichem Augusta Industriale Srl, Milan, Italy
(18-19 October)
IASD
Dr G. Holland, UNILEVER Ltd, Environmental Safety Laboratory,
Sharnbrook, United Kingdom
Dr M. Stalmans, Procter & Gamble ETC, 100 Temselaan,
Strombeek-Bever, Belgium
Secretariat
Dr H.-J. Poremski, Umweltbundesamt, Berlin, Germany (21 October)
Dr E. Smith, International Programme on Chemical Safety, World
Health Organization, Geneva, Switzerland (Secretary)
Dr B. Wittann, Umweltbundesamt, Berlin, Germany (18-20 October)
NOTE TO READERS OF THE CRITERIA MONOGRAPHS
Every effort has been made to present information in the
criteria monographs as accurately as possible without unduly
delaying their publication. In the interest of all users of the
environmental health criteria monographs, readers are requested to
communicate any errors that may have occurred to the Director of the
International Programme on Chemical Safety, World Health
Organization, Geneva, Switzerland, in order that they may be
included in corrigenda, which will appear in subsequent volumes.
* * *
A detailed data profile and a legal file can be obtained from
the International Register of Potentially Toxic Chemicals, Case
Postale 356, 1219 Châtelaine, Geneva, Switzerland (Telephone no.
979 9111).
* * *
This publication was made possible by financial support from the
US Environmental Protection Agency, USA, and from the European
Commission.
ENVIRONMENTAL HEALTH CRITERIA FOR LINEAR ALKYLBENZENE SULFONATES AND
RELATED COMPOUNDS
A WHO Task Group on Environmental Health Criteria for Linear
Alkylbenzene Sulfonates and Related Compounds met at the World
Health Organization, Geneva, on 18-22 October 1993. Dr E. Smith,
IPCS, welcomed the participants on behalf of Dr M. Mercier, Director
of IPCS, and of the three IPCS cooperating organizations (UNEP, ILO,
and WHO). The Group reviewed and revised a draft document and
evaluated the risks for human health and the environment of exposure
to linear alkylbenzene sulfonates, a-olefin sulfonates, and alkyl
sulfonates.
The sections of the first draft on toxicology and human health
were prepared at the National Institute of Health Sciences (NIHS),
Tokyo, Japan, and the sections on the environment at the Institute
of Terrestrial Ecology (ITE), Monks Wood, United Kingdom.
Dr E. Smith of the IPCS Central Unit was responsible for
the scientific content of the monograph and Mrs E. Heseltine,
St Léon-sur-Vézère, France, for the editing.
The authors who contributed to the first draft were:
Dr S. Dobson, ITE, Monks Wood, United Kingdom
Dr R. Hasegawa, NIHS, Tokyo, Japan
Dr Y. Hayashi, NIHS, Tokyo, Japan
Dr K. Hiraga, Public Health Research Laboratory, Tokyo, Japan
Dr P. Howe, ITE, Monks Wood, United Kingdom
Dr Y. Ikeda, NIHS, Tokyo, Japan
Dr Y. Kurokawa, NIHS, Tokyo, Japan
Dr H. Malcolm, ITE, Monks Wood, United Kingdom
Dr A. Matsuoka, NIHS, Tokyo, Japan
Dr K. Morimoto, NIHS, Tokyo, Japan
Dr M. Nakadate, NIHS, Tokyo, Japan
Dr K. Oba, Lion Chemical Corporation, Tokyo, Japan
Dr J. Sekizawa, NIHS, Tokyo, Japan
Dr T. Sohuni, NIHS, Tokyo, Japan
Dr M. Takahashi, NIHS, Tokyo, Japan
Dr R. Takei, Lion Chemical Corporation, Tokyo, Japan
Dr S. Tanaka, NIHS, Tokyo, Japan
Dr S. Tomiyama, Lion Chemical Corporation, Tokyo, Japan
Dr T. Yamaha, NIHS, Tokyo, Japan
Dr S. Yoshikawa, Environmental Research Institute, Kawasaki, Japan
Dr M. Wakabayashi, Water Quality Management Centre, Tokyo, Japan
Dr Y. Watanabe, Central Railway Hospital, Tokyo, Japan
Dr P. Howe, Dr H. Malcolm, and Dr J Sekizawa also contributed to
the second draft.
The efforts of all who helped in the preparation and
finalization of the monograph are gratefully acknowledged.
1. OVERALL SUMMARY, EVALUATION, AND RECOMMENDATIONS
1.1 Identity and analytical methods
Linear alkylbenzene sulfonates (LAS), alpha-olefin sulfonates
(AOS), and alkyl sulfates (AS) are anionic surfactants with
molecules characterized by a hydrophobic and a hydrophilic (polar)
group. Commercial mixtures consist of isomers and homologues of
related compounds, which differ in physicochemical properties,
resulting in formulations for various applications.
LAS, AOS, and AS can be analysed by nonspecific methods. The
assay usually used is one for substances that react with methylene
blue, which responds to any compound containing an anionic and
hydrophobic group. It thus suffers from analytical interference if
used for environmental samples; furthermore, the sensitivity of this
method is about 0.02 mg/litre. Although nonspecific alternatives to
this method have been developed, they are not commonly used.
Specific methods for environmental analysis are available only for
LAS and AS. An improved method based on methylene blue reactivity
and high-performance liquid chromatography (HPLC) is available for
analysis of AOS.
LAS are nonvolatile compounds produced by sulfonation of linear
alkylbenzene. Commercial products are always mixtures of homologues
of different alkyl chain lengths (C10-C13 or C14) and isomers
differing in the phenyl ring positions (2 to 5 phenyl). All of the
homologues and isomers of LAS can be determined in environmental
samples and other matrices by specific analytical methods such as
HPLC, gas chromatography, and gas chromatography-mass spectrometry.
AOS are nonvolatile compounds produced by sulfonation of
alpha-olefins. They are mixtures of two compounds, sodium alkene
sulfonate and hydroxyalkane sulfonate, with alkyl chain lengths of
C14-C18.
AS are nonvolatile compounds produced by sulfation of
oleochemical or petrochemical alcohols. They are mixtures of
homologues with alkyl chain lengths of C10-C18. Specific
analytical methods are being developed for environmental monitoring.
1.2 Sources of human and environmental exposure
LAS, AOS, and AS are used as active ingredients in household and
personal care products and in specialized applications. After use,
such detergent compounds are discharged into the environment in
wastewater.
There is occupational exposure to these compounds. The exposure
of the general human population and of environmental organisms
depends on the application of LAS, AOS, and AS (and other
surfactants), on local sewage treatment practices, and on the
characteristics of the receiving environment.
In 1990, worldwide consumption figures were about 2 million
tonnes of LAS, 86 000 tonnes of AOS, and 289 000 tonnes of AS.
1.3 Environmental concentrations
1.3.1 Linear alkylbenzene sulfonates
Concentrations of LAS have been quantified by means of a
specific, sensitive analytical method in almost every environmental
compartment in which they might be present. The concentrations
decrease progressively in the order wastewater > treated effluent
> surface waters > the sea.
In areas where LAS are the predominant surfactants used, the
concentrations are usually 1-10 mg/litre in wastewater,
0.05-0.1 mg/litre in effluents treated biologically,
0.05-0.6 mg/litre in effluents treated with a percolating filter,
0.005-0.05 mg/litre in surface waters below sewage outfalls (with
concentrations decreasing rapidly to 0.01 mg/litre downstream of the
outfall), < 1-10 mg/kg in river sediments (< 100 mg/kg in highly
polluted sediments near discharge zones), 1-10 g/kg in sewage
sludge, and < 1-5 mg/kg in sludge-amended soils (initially
5-10 mg/kg; - 50 mg/kg have been reported after atypically high
applications of sludge). The concentrations of LAS in estuarine
waters are 0.001-0.01 mg/litre, although higher levels occur where
wastewater is discharged directly. The concentrations in offshore
marine waters are < 0.001-0.002 mg/litre.
It should be noted that the environmental concentrations of LAS
vary widely. This variation is due to differences in analytical
methods, in the characteristics of sampling sites (ranging from
highly polluted areas with inadequate sewage treatment to areas
where sewage undergoes extensive treatment), in season (which can
account for a difference of twofold), and in consumption of LAS.
Environmental monitoring shows that there has been no
accumulation of LAS in environmental compartments over time. The
concentrations in soil do not increase with time but decrease owing
to mineralization. As LAS do not degrade under strictly anaerobic
conditions (to generate methane), it cannot be concluded that they
are mineralized in anaerobic sediments. With current use, the rate
of assimilation of LAS in all receiving environmental compartments
is equal to the rate of input, implying a steady state.
1.3.2 alpha-Olefin sulfonates and alkyl sulfates
Limited data are available on the concentrations of AOS in the
environment owing to the difficulty of analysing them in
environmental samples. Nonspecific colorimetric methods (such as
that based on methylene blue) allow detection of anionic surfactants
in general, but they suffer from analytical interferences and are
not suitable for determining specific concentrations of AOS. A
specific method is being developed for measuring AS in environmental
samples.
Studies conducted in the laboratory indicate that AOS and AS are
mineralized rapidly in all environmental compartments and are
virtually entirely removed from sewage during treatment. The
concentrations in surface water, sediments, soil, estuarine water,
and the marine environment are probably low. The levels of AOS in
river water have been found to be low.
1.4 Environmental transport, distribution, and transformation
At te mperatures below 5-10°C, the biodegradation kinetics of
LAS, AOS, and AS is reduced because of a reduction in microbial
activity.
1.4.1 Linear alkylbenzene sulfonates
The routes by which LAS enter the environment vary among
countries, but the main route is via discharge from sewage treatment
works. When wastewater treatment facilities are absent or
inadequate, sewage may be discharged directly into rivers, lakes,
and the sea. Another route of entry of LAS to the environment is by
the spreading of sewage sludge on agricultural land.
Throughout their passage into the environment, LAS are removed
by a combination of adsorption and primary and ultimate
bio-degradation. LAS are adsorbed onto colloidal surfaces and onto
suspended particles, with measured adsorption coefficients of
40-5200 litres/kg depending on the media and the structure of the
LAS. They biodegrade in surface water (half-life, 1-2 days), aerobic
sediments (1-3 days), and marine and estuarine systems (5-10 days).
During primary sewage treatment, about 25% of LAS (range,
10-40%) are adsorbed onto and removed with waste sludge. They are
not removed during anaerobic sludge digestion but are removed during
aerobic treatment of sludge, with a half-life of about 10 days.
After application of sludge to soil, 90% of LAS are generally
degraded within three months, with a half-life of 5-30 days.
The whole-body concentration factors for LAS range from 100 to
300, for the sum of 14C-LAS and 14C metabolites. Uptake by fish
occurs mainly through the gills, with subsequent distribution to the
liver and gall-bladder after biotransformation. LAS are excreted
rapidly, and there is therefore no evidence that they undergo
biomagnification.
1.4.2 alpha-Olefin sulfonates
Fewer data are available on the environmental transport,
distribution, and transformation of AOS than for LAS. It can be
inferred that AOS are transported into the environment in a manner
similar to that established for LAS, AS and other detergent
surfactants, and the environmental fate of AOS is similar to that of
LAS and AS. It is readily biodegraded under aerobic conditions, and
primary biodegradation is complete within 2-10 days, depending on
the temperature. Limited data are available on the bioaccumulation
of AOS; no bioaccumulation was observed in fish. There are no data
on abiotic degradation.
1.4.3 Alkyl sulfates
AS are transported into the environment by mechanisms similar to
those that operate for LAS and AOS. They are readily biodegradable
under aerobic and anaerobic conditions in the laboratory and under
environmental conditions; primary biodegradation is complete within
2-5 days. The whole-body bioconcentration factor ranges from 2 to 73
and varies with the chain length of alkyl sulfate homologues. AS are
taken up, distributed, biotransformed, and excreted by fish in the
same way as LAS and are not bioconcentrated in aquatic organisms.
1.5 Kinetics
LAS, AOS, and AS are readily absorbed by the gastrointestinal
tract, widely distributed throughout the body, and extensively
metabolized. LAS undergo omega- and ß-oxidation. The parent
compounds and metabolites are excreted mainly through the kidney,
although a proportion of an absorbed dose may be excreted as
metabolites in the faeces by biliary excretion. Only minimal amounts
of LAS, AOS, and AS appear to be absorbed through intact skin,
although prolonged contact may compromise the integrity of the
epidermal barrier, thereby permitting greater absorption; high
concentrations may reduce the time required for penetration.
1.6 Effects on experimental animals and in vitro test systems
The oral LD50 values for sodium salts of LAS were 404-1470
mg/kg body weight in rats and 1259-2300 mg/kg body weight in mice,
suggesting that rats are more sensitive than mice to the toxicity of
LAS. An oral LD50 of 3000 mg/kg body weight was measured for a
sodium salt of AOS in mice. The oral LD50 values of AS in rats
were 1000-4120 mg/kg body weight. LAS, AOS, and AS irritate the skin
and eye.
Minimal effects, including biochemical alterations and
histo-pathological changes in the liver, have been reported in
subchronic studies in which rats were administered LAS in the diet
or drinking-water at concentrations equivalent to doses greater than
120 mg/kg body weight per day. Although ultrastructural changes were
observed in liver cells at lower doses in one study, the changes
appeared to be reversible. Effects were not seen at similar doses in
other studies, but the organs may have been examined more closely in
the initial study. Reproductive effects, including decreased
pregnancy rate and litter loss, have been reported in animals
administered doses > 300 mg/kg per day. Histopathological and
biochemical changes were observed after long-term dermal application
to rats of solutions of > 5% LAS, and after 30 days' application to
the skin of guinea-pigs of 60 mg/kg body weight. Repeated dermal
application of > 0.3% solutions of LAS induced fetotoxic and
reproductive effects, but also induced maternal toxicity. Few data
are available from studies in experimental animals that allow
evaluation of the potential effects of AOS in humans. No effects
were observed in rats administered oral doses of 250 mg/kg body
weight per day chronically, but fetotoxicity was seen in rabbits
administered a maternally toxic dose of 300 mg/kg body weight per
day. Application of AOS to the skin and eyes of experimental animals
induced local effects.
Although the effects of short- and long-term exposure of animals
to AS have been investigated in several studies, most suffered from
inadequate histopathological examination or small group sizes;
furthermore, the highest doses used in the long-term studies did not
produce any toxic effects, so that an NOAEL could not be
established. Effects have, however, been reported consistently in
rats administered AS in the diet or drinking-water at concentrations
equivalent to 200 mg/kg body weight per day or more. Local effects
have been observed on the skin and eyes after topical application of
concentrations of about 0.5% AS or more. Maternally toxic and
fetotoxic effects have been observed at higher concentrations.
Most of the long-term studies are inadequate to evaluate the
carcinogenic potential of LAS, AOS, and AS in experimental animals,
owing to factors such as small numbers of animals, limited numbers
of doses, absence of a maximal tolerated dose, and limited
histo-pathological examination in the majority of studies. In those
studies in which the pathological findings were adequately reported,
maximal tolerated doses were not used, and the doses did not produce
toxic effects. Subject to these limitations, however, the studies in
which animals were administered LAS, AOS, or AS orally gave no
evidence of carcinogenicity; long-term studies in which AOS was
applied by skin painting studies also showed no effect.
On the basis of limited data, these compounds do not appear to
be genotoxic in vivo or in vitro.
1.7 Effects on humans
The results of patch tests show that human skin can tolerate
contact with solutions containing up to 1% LAS, AOS, or AS for 24 h
with only mild irritation reactions. These surfactants caused
delipidation of the skin surface, elution of natural moisturizing
factor, denaturation of the proteins of the outer epidermal layer,
and increased permeability and swelling of the outer layer. Neither
LAS, AOS, nor AS induced skin sensitization in volunteers, and there
is no conclusive evidence that they induce eczema. No serious
injuries or fatalities have been reported following accidental
ingestion of these surfactant by humans.
1.8 Environmental effects
1.8.1 Linear alkylbenzene sulfonates
1.8.1.1 Aquatic environment
LAS have been studied extensively both in the laboratory (short-
and long-term studies) and under more realistic conditions (micro-
and mesocosm and field studies). In general, a decrease in alkyl
chain length or greater internalization of the phenyl group is
accompanied by a decrease in toxicity. Observations in fish and
Daphnia indicate that a decrease in chain length of one unit (e.g.
C12 to C11) results in an approximately twofold decrease in
toxicity.
The results of laboratory tests are as follows:
-- Microorganisms: The results are highly variable owing to
the use of a variety of test systems (e.g. inhibition of activated
sludge; mixed cultures and individual species). The EC50 values
range from 0.5 mg/litre (single species) to > 1000 mg/litre. For
microorganisms, there is no linear relationship between chain length
and toxicity.
-- Aquatic plants: The results are highly species dependent.
For freshwater organisms, the EC50 values are 10-235 mg/litre
(C10-C14) in green algae, 5-56 mg/litre (C11.1-C13) in blue
algae, 1.4-50 mg/litre (C11.6-C13) in diatoms, and
2.7-4.9 mg/litre (C11.8) in macrophytes; marine algae appear to be
even more sensitive. In algae, there is probably no linear
relationship between chain length and toxicity.
-- Invertebrates: The acute L(E)C50 values for at least 22
freshwater species are 4.6-200 mg/litre (chain length not specified;
C13) for molluscs; 0.12-27 mg/litre (not specified; C11.2-C18)
for crustaceans; 1.7-16 mg/litre (not specified; C11.8) for worms,
and 1.4-270 mg/litre (C10-C15) for insects. The chronic L(E)C50
values are 2.2 mg/litre (C11.8) for insects and 1.1-2.3 mg/litre
(C11.8-C13) for crustaceans. The chronic no-observed-effect
concentration (NOEC; based on lethality or reproductive effects) is
0.2-10 mg/litre (not specified; C11.8) for crusta-ceans. Marine
invertebrates appear to be more sensitive, with LC50 values of 1
to >100 mg/litre (almost all C12) for 13 species, and NOECs of
0.025-0.4 mg/litre (not specified for all tests) for seven species
tested.
-- Fish: The acute LC50 values are 0.1-125 mg/litre
(C8-C15) for 21 freshwater species; the chronic L(E)C50 values
are 2.4 and 11 mg/litre (not specified; C11.7) for two species;
and the NOECs are 0.11-8.4 to 1.8 mg/litre (not specified;
C11.2-C13) for two species. Again, marine fish appear to be more
sensitive, with acute LC50 values of 0.05-7 mg/litre (not
specified; C11.7) for six species and chronic LC50 values of
0.01-1 mg/litre (not specified) for two species. In most of the
reports, the chain length was not reported. An NOEC of <
0.02 mg/litre (C12) was reported for marine species.
The average chain length of products commonly used commercially
is C12. Compounds of many different chain lengths have been tested
in Daphnia magna and fish, but the length tested in other
freshwater organisms has usually been C11.8. The typical acute
L(E)C50 values for C12 LAS are 3-6 mg/litre in Daphnia magna
and 2-15 mg/litre in freshwater fish, and the typical chronic NOECs
are 1.2-3.2 mg/litre for Daphnia and 0.48-0.9 mg/litre for
freshwater fish. The typical acute LC50 values for LAS of this
chain length in marine fish are < 1-6.7 mg/litre.
Saltwater organisms, especially invertebrates, appear to be more
sensitive to LAS than freshwater organisms. In invertebrates, the
sequestering action of LAS on calcium may affect the availability of
this ion for morphogenesis. LAS have a general effect on ion
transport. Biodegradation products and by-products of LAS are 10-100
times less toxic than the parent compounds.
The results obtained under more realistic conditions are as
follows: LAS have been tested in all freshwater tests at several
trophic levels, including enclosures in lakes (lower organisms),
model ecosystems (sediment and water systems), rivers below and
above the outfall of wastewater treatment plants, and in
experimental streams. C12 LAS were used in almost all cases. Algae
appear to be more sensitive in summer than in winter, as the 3-h
EC50 values were 0.2-8.1 mg/litre after photosynthesis, whereas in
model ecosystems no effects were seen on the relative abundance of
algal communities at 0.35 mg/litre. The no-effect levels in these
studies were 0.24-5 mg/litre, depending on the organism and
parameter tested. These results agree fairly well with those of
laboratory tests.
1.8.1.2 Terrestrial environment
Information is available for plants and earthworms. The
NOECs for seven plant species tested in nutrient solutions are
< 10-20 mg/litre; that for three species tested in soils, based
on growth, was 100 mg/kg (C10-C13). The 14-day LC50 for earthworms
was > 1000 mg/kg.
1.8.1.3 Birds
One study of chickens treated in the diet resulted in an NOEC
(based on egg quality) of > 200 mg/kg.
1.8.2 alpha-Olefin sulfonates
There are limited data on the effects of AOS on aquatic and
terrestrial organisms.
1.8.2.1 Aquatic environment
Only the results of laboratory tests are available:
-- Algae: EC50 values of > 20-65 mg/litre (C16-C18)
have been reported for green algae.
-- Invertebrates: LC50 values of 19 and 26 mg/litre (C16-C18)
have been reported for Daphnia.
-- Fish: The acute LC50 values are 0.3-6.8 mg/litre (C12-C18)
for nine species of fish. On the basis of short-term studies in
brown trout (Salmo trutta), golden orfe (Idus melanotus), and
harlequin fish (Rasbora heteromorpha), it can be concluded that
the toxicity of C14-C16 compounds is about five times lower than
that of C16-C18, with LC50 values (all measured concentrations)
of 0.5-3.1 (C16-C18) and 2.5-5.0 mg/litre (C14-C16). Two
long-term studies in rainbow trout showed that growth is the most
sensitive parameter, resulting in an EC50 of 0.35 mg/litre. In a
marine fish, the grey mullet (Mugal cephalus), the 96-h LC50 value
was 0.70 mg/litre.
1.8.2.2 Terrestrial environment
One study of plants in nutrient solutions showed NOECs of
32-56 mg/litre. In a study of chickens treated in the diet, an NOEC
(based on egg quality) of > 200 mg/kg was reported.
1.8.3 Alkyl sulfates
1.8.3.1 Aquatic environment
AS have been studied in short- and long-term studies and in one
study under more realistic conditions. Their toxicity is again
dependent on the alkyl chain length; no information was available on
any difference in toxicity between linear and branched AS.
The results of the laboratory tests are as follows:
-- Microorganisms: The EC50 values in a marine community
were 2.1-4.1 mg/litre (C12). The NOECs in Pseudomonas putida were
35-550 mg/litre (C16-C18).
-- Aquatic plants: The EC50 values were > 20-65 mg/litre
(C12-C13) in green algae and 18-43 mg/litre (C12) in
macrophytes. The NOECs were 14-26 mg/litre (C12-C16/C18) in
green algae.
-- Invertebrates: The LC50 and EC50 values were 4-140 mg/litre
(C12/C15-C16/C18) in freshwater species and 1.7-56 mg/litre
(all C12) in marine species. The chronic NOEC in Daphnia magna was
16.5 mg/litre (C16/C18) and those in marine species were
0.29-0.73 mg/litre (chain length not specified).
-- Fish: The LC50 values were 0.5-5.1 mg/litre (not
specified; C12-C16) in freshwater species and 6.4-16 mg/litre
(all C12) in marine species. No long-term studies were available.
It should be noted that many of these studies were carried out
under static conditions. As AS are readily biodegradable, their
toxicity may have been underestimated. In a 48-h study with Oryzias
latipes, the LC50 values were 46, 2.5 and 0.61 mg/litre
(measured concentrations) for C12, C14, and C16 compounds,
respectively. This and other studies indicate that toxicity differs
by a factor of five for two units of chain length. In a flow-through
biocenosis study with compounds of C16-C18, an NOEC of
0.55 mg/litre was observed.
1.8.3.2 Terrestrial environment
NOEC values of > 1000 mg/kg (C16-C18) were reported for
earthworms and turnips.
1.9 Human health risk evaluation
LAS are the most widely used surfactants in detergents and
cleaning products; AOS and AS are also used in detergents and
personal care products. The primary route of human exposure is,
therefore, through dermal contact. Minor amounts of LAS, AOS, and AS
may be ingested in drinking-water and as a result of residues on
utensils and food. Although limited information is available, the
daily intake of LAS via these media can be estimated to be about
5 mg/person. Occupational exposure to LAS, AOS, and AS may occur
during the formulation of various products, but no data are
available on the effects in humans of chronic exposure to these
compounds.
LAS, AOS, and AS can irritate the skin after repeated or
prolonged dermal contact with concentrations similar to those found
in undiluted products. In guinea-pigs, AOS can induce skin
sensitization when the level of gamma-unsaturated sultone exceeds
about 10 ppm.
The available long-term studies in experimental animals are
inadequate to evaluate the carcinogenic potential of LAS, AOS, and
AS, owing to factors such as study design, use of small numbers of
animals, testing of insufficient doses, and limited
histopathological examination. In the limited studies available in
which animals were administered LAS, AOS, or AS orally, there was no
evidence of carcinogenicity; the results of long-term studies in
which AOS were administered by skin painting were also negative.
These compounds do not appear to be genotoxic in vivo or
in vitro, although few studies have been reported.
Minimal effects, including biochemical alterations and
histopathological changes in the liver, have been reported in
subchronic studies of rats administered LAS in the diet or
drinking-water at concentrations equivalent to a dose of about
120 mg/kg body weight per day, although no effects were observed in
studies in which animals were exposed to higher doses for longer
periods. Dermal application of LAS caused both systemic toxicity and
local effects.
The average daily intake of LAS by the general population, on
the basis of limited estimates of exposure via drinking-water,
utensils, and food, is probably much lower (about three orders of
magnitude) than the levels shown to induce minor effects in
experimental animals.
The effects of AOS in humans observed in the few studies
available are similar to those reported in animals exposed to LAS.
As insufficient data are available to estimate the average daily
intake of AOS by the general population and on the levels that
induce effects in humans and animals, it is not possible to evaluate
with confidence whether exposure to AOS in the environment presents
a risk to human health. The levels of AOS in media to which humans
may be exposed are likely to be lower than those of LAS, however, as
AOS are used less.
Effects have been reported consistently in a few, limited
studies in rats administered AS in the diet or drinking-water at
concentrations equivalent to doses of 200 mg/kg body weight per day
or more. Local effects on the skin and eyes have been observed after
repeated or prolonged topical application. The available data are
insufficient to estimate the average daily intake of AS by the
general population. Since AS surfactants are not used as extensively
as those containing LAS, however, intake of AS is likely to be at
least three orders of magnitude lower than the doses shown to induce
effects in animals.
1.10 Evaluation of effects on the environment
LAS, AS, and AOS are used in large quantities and are released
into the environment via wastewater. Risk assessment requires
comparison of exposure concentrations with concentrations that cause
no adverse effects, and this can be done for several environmental
compartments. For anionic surfactants in general, the most important
compartments are sewage water treatment plants, surface waters,
sediment- and sludge-amended soils, and estuarine and marine
environments. Both biodegradation (primary and ultimate) and
adsorption occur, resulting in decreased environmental
concentrations and bioavailability. Reduction in chain length and
loss of the parent structure both result in compounds that are less
toxic than the parent compound. It is important that these
considerations be taken into account when the results of laboratory
tests are compared with potential effects on the environment.
Furthermore, in assessing the risk associated with environmental
exposure to these three anionic compounds, comparisons should be
made with the results of tests for toxicity of compounds of the same
chain length.
The effects of LAS on aquatic organisms have been tested
extensively. In laboratory tests in freshwater, fish appeared to be
the most sensitive species; the NOEC for fathead minnow was about
0.5 mg/litre (C12), and these results were confirmed in tests
under more realistic conditions. Differences have been observed
among phyto-plankton: in acute 3-h assays on phytoplankton, the
EC50 values were 0.2-0.1 mg/litre (C12-C13), whereas no
effects on relative abundance were found in other tests at
0.24 mg/litre (C11.8). Marine species appeared to be slightly more
sensitive than most other taxonomic groups.
A broad range of concentrations of all three anionic compounds
occurs in the environment, as shown by extensive measurements of
LAS. Owing to this broad range, no generally applicable
environmental risk assessment can be made for these compounds. A
risk assessment must involve appropriate understanding of the
exposure and effect concentrations in the ecosystem of interest.
Accurate data on exposure to AS and AOS are needed before an
environmental risk assessment can be made. Models are therefore
being used to assess exposure concentrations in the receiving
environmental compartments. Data on the toxicity of AS and AOS to
aquatic organisms, especially after chronic exposure to stable
concentrations, are relatively scarce. The available data show that
the toxicity of AOS and AS is similar to that of other anionic
surfactants.
Saltwater organisms appear to be more sensitive than freshwater
organisms to these compounds; however, their concentrations are
lower in seawater, except near wastewater outlets. The fate and
effects of these compounds in sewage in seawater have not been
investigated in detail.
For an evaluation of the environmental safety of surfactants
such as LAS, AOS, and AS, actual environmental concentrations must
be compared with no-effect concentrations. Research requirements are
determined not only by the intrinsic properties of a chemical but
also by its pattern or trend of consumption. As these can vary
considerably among geographic areas, assessment and evaluation must
be carried out regionally.
1.11 Recommendations for protection of human health and
the environment
1. As exposure to dusts may occur in the workplace (during
processing and formulation), standard occupational hygiene practices
should be used to ensure protection of workers' health.
2. The composition of formulations for consumer and industrial use
should be designed to avoid hazard, particularly for formulations
that are used for cleaning or laundering by hand.
3. Environmental exposure and effects should be appropriately
monitored to provide early indications of any overloading of
relevant environmental compartments.
1.12 Recommendations for further research
Human health
1. Since the skin is the primary route of human exposure to LAS,
AOS, and AS and since no adequate long-term studies of dermal
toxicity or carcinogenicity in experimental animals are available,
it is recommended that suitably designed long-term studies in which
these compounds are applied dermally be conducted.
2. In view of the lack of definitive data on the genotoxicity of
AOS and AS, additional studies should be performed in vivo and
in vitro.
3. In view of the inadequacies of the available studies on
reproductive and developmental toxicity, definitive studies should
be carried out in laboratory animals to obtain data on the effects
and on the effect and no-effect levels of LAS, AOS, and AS.
4. As exposure to LAS, AOS, and AS is not adequately defined, the
exposure of the general population should be monitored, particularly
when these surfactants are used for cleaning and laundering by hand.
5. Since LAS, AOS, and AS may enhance the transport of other
chemicals in environmental media and modulate their bioavailability
and toxicity in surface waters, river sediments, and soils to which
humans may be exposed, interactions with other environmental
chemicals and the consequences for humans should be investigated.
Environmental safety
6. Additional studies should be carried out on the mechanisms of
adsorption and desorption of AOS and AS. Studies should also be done
on the partitioning of LAS, AOS, and AS between dissolved and
suspended colloidal particles in water. Mathematical models of
sorption coefficients should be developed and validated on the basis
of physical-chemical parameters.
7. Studies of the biodegradation of AOS and AS in sludge-amended
soils and river sediments should be carried out when exposure
occurs. Studies in river sediments (aerobic and anaerobic zones)
should be performed downstream of treated and untreated wastewater
and sewage outfalls.
8. Environmental concentrations of LAS, AOS, and AS should be
monitored regionally and nationally in order to obtain information
on exposure. Analytical methods should be developed for detecting
low levels of AOS and AS in relevant environmental compartments.
9. National databases should be developed on the concentrations of
LAS, AOS, and AS in wastewater and rivers and on the types,
efficiency, and location of wastewater treatment plants, in order to
facilitate an assessment of the impact of discharges of these
surfactants to the environment.
10. Long-term studies of the toxicity of AOS and AS to fish
(freshwater and marine) and aquatic invertebrates should be
conducted in order to establish the relative sensitivity of these
species.
A. Linear alkylbenzene sulfonates and their salts
A1. SUMMARY
See Overall Summary, Evaluation, and Recommendations (pp. 7-21).
A2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND
ANALYTICAL METHODS
A2.1 Identity (sodium salt)
Chemical formula: CnH2n-1O3S Na ( n: 16-20) (for
current commercial products)
Chemical structure:
j,k: integers ( j + k = 7-11)
Common name: Sodium linear alkylbenzenesulfonate
Common synonyms: LAS, LAS sodium salt, linear
alkylbenzene-sulfonic acid sodium salt,
linear dodecyl-benzenesulfonic acid
sodium salt, sodium straight chain
alkylbenzenesulfonate
CAS Registry number: 68411-30-3 (LAS sodium salt, C10-13
alkyl)
Common trade names: Ablusol DBC, Agrilan WP, Alkasurf CA,
Arylan, Atlas G-3300B, Atlox, Biosoft,
Berol, Calsoft, Demelan CB-30, Elecut
S-507, Elfan, Emulphor ECB, Emulsogen
Brands, Gardilene, Hexaryl, Idet,
Kllen, Lutopon SN, Manro, Marlopon,
Marlon A, Nacconol 90 F, Nansa HS 80,
Nansa Lutersit, Neopelex, Sandozin AM,
Sipex, Sulfamin, Sulframin, Surfax 495,
Teepol, Tersapol, Tersaryl, Ufaryl DL
80P, Witconate (McCutcheon, 1993)
Abbreviations: LAS, LAS-Na
Specification: LAS are anionic surfactants which were
introduced in the 1960s as more
biodegradable replacements for highly
branched alkyl-benzene sulfonates. LAS
are produced by sulfonation of linear
alkylbenzene (LAB) with sulfur trioxide
(SO3), usually on a falling film
reactor or with oleum in batch
reactors. The corresponding sulfonic
acid is subsequently neutralized with
an alkali such as caustic soda. The
hydrocarbon intermediate, LAB, is
currently produced mainly by alkylation
of benzene with n-olefins or
n-chloroparaffins using hydrogen
fluoride (HF) or aluminium chloride
(AlCl3) as a catalyst, and the LAS
derivatives are thus generally referred
to in that context (Cavalli et al.,
1993a). Currently, 74% of world
production of LAB is via HF and 26% via
AlCl3 (Berna et al., 1993a).
LAS are a mixture of homologues and phenyl positional isomers,
each containing an aromatic ring sulfonated at the para position
and attached to a linear alkyl chain of C10-C14 (in Europe,
predominantly C10-C13) at any position except the terminal one.
The product is generally used in detergents in the form of the
sodium salt.
Some of the typical characteristics of LAS, including the
distribution of alkyl chain lengths and the positions of the phenyl
rings in the two types of LAS used in laundry detergents, are shown
in the box below. The United States Toxic Substances Control Act
inventory lists LAS homologues with chain lengths up to C18
(Tables 1 and 2), but these products are not currently used for
commercial purposes.
A2.2 Physical and chemical properties
The properties of LAS differ greatly depending on the alkyl
chain length. Table 3 shows the Krafft points (temperature at which
1 g of LAS dissolve in 100 ml of water) and the relative critical
micelle concentrations of the single homologues.
Typical characteristics of linear alkylbenzene sulfonates used in laundry
detergents:
Appearance (commercial product): White paste (containing water)
Average length of alkyl carbon chain: 11.8
Average relative molecular mass: 342
Unsulfonated matter: 1-2%
Alkyl chain distribution:
C10 10-15%
C11 25-35%
C12 25-35%
C13 15-30%
C14 0-5%
Phenyl ring position LAS (LAB-HFa) LAS (LAB-AlCl3b)
2-phenyl 18 28
3-phenyl 16 19
4-phenyl 17 17
5-phenyl 24 18
6-phenyl 25 18
From Cavalli et al. (1993a)
a Hydrofluoric acid-catalysed process
b Aluminium chloride-catalysed process
Table 1. Mixtures of linear alkylbenzene sulfonates and their salts found in the
United States Toxic Substances Control Act inventory
Generic benzene- CAS number
sulfonic acid groups
Acid Salts
(C10-13)Alkyl-a 68411-30-3 (sodium salt)
(C10-16)Alkyl- 68584-22-5 68584-23-6 (calcium salt)
68584-26-9 (magnesium salt)
68584-27-0 (potassium salt)
Mono (C6-12)alkyl- 68608-87-7 (sodium salt)
Mono(C7-17)alkyl- 68953-91-3 (calcium salt)
68953-94-6 (potassium salt)
Mono(C9-12)alkyl- 68953-95-7 (sodium salt)
Mono(C10-16)alkyl- 68910-31-6 (ammonium salt)
68081-81-2 (sodium salt)
Mono(C12-18)alkyl- 68648-97-5 (potassium salt)
a There may be more than one alkyl substituent per benzene ring (United
States Environmental Protection Agency, 1981).
Table 2. Individual linear alkylbenzene sulfonates (LAS) found in the United States Toxic Substances Control Act inventory
Parent sulfonic acid Empirical CAS Registry number
(abbreviation) formula
Acids Sodium salts Other salts
Dodecylbenzene C16H26O3S 1322-98-1 1322-98-1
(C10 LAS) (140-60-3)a (2627-06-7)a
Undecylbenzene C17H28O3S 50854-94-9 27636-75-5 NH4 salt, 61931-75-7
(C11 LAS)
Dodecylbenzene C18H30O3S 27176-87-0 25155-30-0 Al salt, 29756-98-7; NH4 salt, 1331-61-9;
(C12 LAS) (2211-98-5)a Ca salt, 26264-06-2; K salt, 27177-77-1;
(68628-60-4)b also numerous salts with alkyl amines
(18777-54-3)c
Tridecylbenzene C19H32O3S 25496-01-9 26248-24-8 Also salts with alkyl amines
(C13 LAS)
Tetradecylbenzene C20H34O3S 30776-59-1 28348-61-0
(C14 LAS) (47377-10-2)a (1797-33-7)a
Pentadecylbenzene C21H36O3S 61215-89-2 K salt, 64716-02-5
(C15 LAS)
Hexadecylbenzene C22H38O3S (16722-32-0)a K salt, 64716-00-3
(C16 LAS)
Heptadecylbenzene C23H40O3S 39735-13-2
(C17 LAS)
From United States Environmental Protection Agency (1981)
a Specifies para substitution
b Specifies para substitution at second position on alkyl chain
Table 3. Relationship between alkyl chain length, Krafft point,
and critical micelle concentration (CMC) of linear
alkylbenzene sulfonates
Alkyl chain length Krafft point (°C) CMC × 10-3 (25°C)
10 -1 5.8
12 3 1.1
14 8 0.24
15 - 0.11
16 13 -
From Ohki & Tokiwa (1970)
The solubility of surfactants in water, defined as the
concentration of dissolved molecules in equilibrium with a
crystalline surfactant phase, increases with rising temperature. For
surfactants, a distinct, sharp bend (break point) is observed in the
solubility/temperature curve. The steep rise in solubility above the
sharp bend is caused by micelle formation. The point of intersection
of the solubility and critical micelle curves plotted as a function
of temperature is referred to as the Krafft point, which is a triple
point at which surfactant molecules coexist as monomers, micelles,
and hydrated solids. The temperature corresponding to the Krafft
point is called the Krafft temperature. Above the Krafft temperature
and critical micelle concentration, a micellar solution is formed
and higher than aqueous solubility may be obtained.
As commercial LAS are a mixture of homologues and phenyl-
positional isomers, their properties may differ. Even some products
with the same alkyl chain distribution (same average carbon number)
have different properties, depending on the 2-phenyl isomer content.
The solubility in water of commercial LAS used for detergents
(average alkyl carbon length, 11.8), for example, which is important
for liquid formulations, is typically about 25% at 25°C for LAS (LAB
via HF) and about 38% at 25°C for LAS (LAB via AlCl3) (Cavalli et
al., 1993a).
As LAS are anionic surfactants, they lower the surface tension
of water so that it can wet and penetrate fabrics more easily to
loosen and remove soils and stains. Micelles, which are formed at
low concentrations, solubilize oil and stains effectively (Ohki &
Tokiwa, 1969). Other important properties of LAS are detergency,
foaming, sensitivity to Ca and Mg ions, wetting, and surface
tension, which reach their optimal values generally when the alkyl
chain length is about C12 (Yamane et al., 1970).
A physico-chemical property often used in environmental
modelling is the octanol-water partition coefficient (Kow).
Although it is impossible to measure the Kow for surface-active
compounds like LAS, it can be calculated. Roberts (1989) modified
the fragment method of Leo & Hansch (1979) in order to take the
branching of position into account. He thus defined a function, log
( CP + 1), where CP is found by pairing off carbon atoms along
the two branches up to the terminus of the shorter branch. (In the
case of LAS, CP is the carbon number of the shorter of the
integers j and k noted in section 2.1.) This gave the formula:
log Kow = ALK-1.44 log ( CP + 1),
where ALK is log Kow calculated without a branch factor.
In order to calculate log Kow for multicomponent materials
like LAS, the calculated Kow for each component is multiplied by
the mole fraction of the corresponding component, the products are
summed, and the logarithm is calculated to give log WAK ( WA,
weighted average).
A2.3 Analysis
A2.3.1 Isolation
A number of analytical methods are available for the
determination of LAS in water, but the primary method is assay as
methylene blue-active substances (MBAS). The methylene blue reaction
responds to any compound containing an anionic centre and a
hydrophobic centre, because such compounds tend to form an
extractable ion pair when they combine with cationic dyes such as
methylene blue; as only the oxidized form is blue, many positive
interferences may occur. Negative interference in MBAS analysis is
seen in the presence of cationic substances such as proteins and
amines (Swisher, 1970, 1987). Therefore, isolation of LAS from a
sample is one of the most important aspects of their analysis. Most
analytical methods include appropriate procedures for isolation.
A2.3.2 Analytical methods
The analytical methods available for determining LAS in water
include nonspecific methods, involving colorimetric, fluorimetric,
and atomic adsorption techniques, and specific methods involving
techniques such as high-performance liquid chromatography (HPLC),
gas chromatography (GC) and GC-mass spectrometry (MS).
A2.3.2.1 Nonspecific methods
The simplest procedure for the determination of LAS in aqueous
solution is a two-phase titration method. LAS are titrated in a
mixed aqueous chloroform medium with a standard solution of a
cationic reagent, such as benzethonium chloride (Hyamine 1622), and
a small amount of indicator, such as a mixture of dimidium bromide
and acid blue. The end-point is determined by a change in the colour
of the organic solvent (ISO 2271, 1972).
The main nonspecific analytical method used is assay for MBAS,
described above. Colorimetric techniques are routinely used to
determine low concentrations of anionic surfactants, including LAS,
in aqueous samples and have been used extensively in testing and
environmental monitoring of these materials. The colorimetric
methods have the same common analytical basis, that is, formation of
solvent extractable compounds between the anionic surfactant and an
intensely coloured cationic species. The most commonly used cationic
reagent for this purpose is methylene blue (Swisher, 1970, 1987).
The same principle has been used as the basis of many other
procedures for the determination of anionic surfactants.
It has been shown or predicted that organic sulfates,
sulfonates, carboxylates, phenols, and even simple inorganic anions
such as cyanide, nitrate, thiocyanate, and sulfate can be methylene
blue-reactive (Swisher, 1970, 1987). The negative interferences that
can occur as a result of direct competition of other 'cationic'
materials are generally considered to be less important than
positive interferences, and the entities detected by the analysis
are correctly referred to as MBAS.
The procedure developed by Longwell & Maniece (1955) and the
improved version of Abbott (1962) are considered to be the best
methods for the determination of MBAS in aqueous samples. The
sensitivity of these procedures is such that levels of
0.01-0.02 mg/litre MBAS can be determined.
The MBAS response can be used as an acceptable overestimate of
the synthetic anionics present in domestic wastewaters, but these
materials may comprise only a small proportion of the total MBAS in
surface waters (Waters & Garrigan, 1983; Matthijs & De Henau, 1987).
Berna et al. (1991) found that LAS contributed 75% of the MBAS in
integrated sewage and 50% in treated water. Direct methylene blue
analysis of extracts derived from sludge, sediment, and soil
invariably leads to highly inflated estimates of LAS (Matthijs & De
Henau, 1987). Numerous attempts have been made to improve the
specificity of methylene blue analysis, by using a variety of
separation steps before the usual colorimetric estimation. Such
indirect procedures are usually lengthy, difficult, and still
susceptible to interference. A number of analytical methods for the
determination of LAS involving extraction and methylene blue are
summarized in Table 4.
Table 4. Analytical methods for anionic surfactants in environmental water using methylene
blue and extraction
Method Isolation method/ Limit of Interference Reference
procedure detection
(mg/litre)
Absorption Extract LAS in water 50-300 Urea, Jones (1945)
photometry into chloroform as thiocyanate,
ion-pair with MB; measure chloride
absorption of chloroform
solution at 650 nm
Extract from alkaline 10-100 As above Longwell &
solution, wash with Maniece
cidic MB (1955)
Remove impurities 0.1-1 As above Abbot (1962)
from MBreagent by
chloroformextraction
Remove MBAS by TLC 0.1-1 Oba & Yoshida
(1965)
Remove MBAS on Takeshita &
polymer bead column Yoshida
(1975)
Remove MBAS on ion 0.02 Yasuda
exchange column (1980)
UV absorption Re-extract LAS into 1 Uchiyama
photometry water; measure UV (1977)
absorption at 222 nm
Table 4 (contd)
Method Isolation method/ Limit of Interference Reference
procedure detection
(mg/litre)
Infra-red Use to reduce 1000 Ambe &
spectometry interference from MBAS Hanya
(1972)
Gas Convert into fluorine 0.02 Tsukioka &
chromatography derivative; measure Murakami
by ECD (1983)
HPLC Remove MB by cation 0.1 Hashimoto et
exchange, HPLC al. (1976)
Remove MB by anion 0.02 Saito et al.
exchange, HPLC (1982)
LAS, linear alkylbenzene sulfonates; MB, methylene blue; MBAS, methylene blue-active
substances; TLC, thin-layer chromatography; UV, ultraviolet radiation; ECD, electron
capture detection; HPLC, high-performance liquid chromatography
Many other cationic dyes and metal chelates have been used as
colorimetric (and fluorimetric) reagents for the determination of
anionic surfactants, including LAS. Use of the cationic metal
chelates has also led to the development of sensitive atomic
absorption methods for indirect determination of anionic surfactants
in fresh, estuarine, and marine waters. Although these alternative
systems may offer some advantages over the methylene blue cation
method, they cannot match the wide experience gained with methylene
blue analysis. Some examples of analytical methods based on the use
of alternative cationic reagents are shown in Table 5.
A2.3.2.2 Specific methods
Good progress has been made towards developing methods for the
specific determination of the many homologues and phenyl-positional
isomers of LAS in almost all laboratory and environmental matrices
(liquid and solid) at concentrations down to micrograms per litre.
High-resolution GC techniques have allowed determination of all the
major components of LAS (homologues and phenyl-positional isomers)
in environmental samples. Waters & Garrigan (1983) and Osburn (1986)
reported improved microdesulfonation-GC procedures for the
determination of LAS in both liquid and solid matrices.
Derivatization techniques offer an alternative approach to
desulfonation for increasing the volatility of LAS for GC (or GC-MS)
analysis (Hon-nami & Hanya, 1980a; McEvoy & Giger, 1986; Trehy et
al., 1990. The GC-MS technique was also applied, after ion-pair,
supercritical fluid extraction and derivatization, to five sewage
sludges, and the LAS were found to occur at 3.83-7.51 g/kg on a
daily basis (Field et al., 1992). These GC procedures, however,
involve extensive sample pre-treatment and depend on conversion of
the isolated LAS into a suitably volatile form for GC determination;
they are therefore time-consuming.
HPLC offers a more convenient means for determining homologues
of LAS in all types of environmental matrices routinely. Several
researchers have reported HPLC procedures for LAS which involve
trace enrichment of the surfactant as the first step (Kikuchi et
al., 1986; Matthijs & De Henau, 1987; Castles et al., 1989; Di
Corcia et al., 1991). Takita & Oba (1985) developed a modified
analytical method based on MBAS-HPLC measurement. Further HPLC
methods, some requiring no sample preparation, are listed in
Table 6.
Table 5. Analytical methods involving reagents other than methylene blue
Method Isolation method/ Limit of Interference Reference
procedure detection
(mg/litre)
Absorption 1-Methyl-4-(4-diethyl- 0.04 Fe[III] Higuchi et al.
photometry aminophenylazo)pyridinium (1982)
iodide; measure
chloroform solution at
564 nm
Bis[2-(5-chloro-2- 0.06 Taguchi et al.
pyridylazo)-5-diethyl- (1981);
aminophenolato]Co Kobayashi
[III] chloride; measure et al. (1986)
benzene solution at
560 nm
Ethylviolet; measure 0.01 Motomizu et
benzene or toluene al. (1982);
solution at 540 nm Yamamoto &
Motomizu (1987)
Atomic Bis[2-(5-chloro-2-
absorption pyridylazo)-5- 1 × 10-3 Hydro- Adachi &
spectrometry diethylaminophenolato] chlorite Kobayashi
Co [III] chloride; measure ion (1982)
Co by atomic absorption
spectrometry
Potassium dibenzo- 0.05 Alkali, Nakamura et al.
18-crown-6; measure K alkaline (1983)
earth
metals
Table 5 (contd)
Method Isolation method/ Limit of Interference Reference
procedure detection
(mg/litre)
Atomic Cu[II] ethylenediamine 0.03 × 10-3 Gagnon (1979);
absorption derivatives; measure Cu Sawada et al.
spectrometry (1983)
Absorption Bis(ethylenediamine)Cu; 5 × 10-3 Rama Bhat
photometry determine Cu after et al. (1980)
addition of 1-(2-
pyridylazo)-2-naphthol
at 560 nm
GC-MS Extract solid phase on 1 × 10-3 Trehy et al.
C8 column; derivatize (1990)
LAS with sulfonyl
chloride for GC-MS
LAS, linear alkylbenzene sulfonates; GC-MS, gas chromatography-mass spectrometry
Table 6. Analytical methods for linear alkylbenzene sulfonates (LAS) by specific analysees
Extraction method Analytical conditions Limit of detection Reference
(mg/litre)
Recover LAS on column Column, silica gel, mobile phase 0.02-0.03 Takano et al. (1975)
chromatograph packed hexane:ethanol containing
with polymer beads sulfuric acid; UV at 225 nm
Extract LAS with methylisobutyl Column, ODS; mobile phase, 0.05 Matsueda et al.
ketone ethanol:water; UV at 225 nm (1982)
Recover LAS by ionexchange Column, cyanopropyl-modified silica; 0.04 Saito et al.
column mobile phase, ethanol:water; (1982)
chromatography UV at 225 nm
Direct analysis Column, ODS; mobile phase, methanol: 0.1 Nakae et al.
water with sodium perchlorate; (1980)
fluorescence detector capable of
determining alkyl homologue distribution
Extract LAS using Column, ODS; mobile phase, acetonitrile: 0.1 × 10-3 Kikuchi et al.
mini-column water with sodium perchlorate; (1986)
fluorescence detector
Concentrate LAS using Column, ODS; mobile phase, acetonitrile: 3 × 10-3 Takami et al.
mini-cartridge column water with sodium perchlorate; (1987)
connected in sequence with fluorescence detector
HPLC system
Table 6 (contd)
Extraction method Analytical conditions Limit of detection Reference
(mg/litre)
Extract LAS with methylisobutyl Column, ODS; mobile phase, acetonitrile: NR Inaba & Amano
ketone water (gradient elution to sharpen peak) (1988)
with sodium perchlorate; UV at 222 nm
Extract from solids with Column, octyl-modified silica; 0.8 Marcomini &
methanol on Soxhlet mobile phase, 2-propanol:water: (injected Giger
acetonitrile (gradient elution) weight) (1987)
with sodium perchlorate; fluorescence
detector
Two-step solid phase Column, C1 Sphesorb; mobile phase, 7 × 10-3 Castles et al. (1989)
extraction with C2 and THF:water with sodium perchlorate;
SAX cartridges fluorescence detector
Extract LAS using Column, C8-DB (Supelco); mobile 0.8 × 10-3 Di Corcia et al.
Carbopack B (graphitized phase, methanol:water with sodium (1991)
carbon black) cartridge perchlorate; fluorescence detector
Concentrate LAS on Column, Wakosil 5C4; mobile phase, 10 × 10-3 Yokoyama & Sato
anion-exchange pre-column acetonitrile:water with sodium (1991)
connected to HPLC system perchlorate; UV at 220 nm
Table 6 (contd)
Extraction method Analytical conditions Limit of detection Reference
(mg/litre)
Ion-pair extraction under Column, capillary gas chromatography, NR Field et al.
SFE conditions using 20 m; mass spectrometry with electron (1992)
tetralhyl-ammonium ion impact ionization operating in
pair reagents, coupled with selected ion mode
ion-pair derivatization
Solid-phase extraction for HPLC column, Bandapat C18 10 × 10-3 Matthijs & De Henau
purification and gradient elution water:acetonitrile (water phase) (1987)
concentration and 0.15 mol/litre NaClOn 0.1 (solid phase)
UV, ultraviolet spectrometry; ODS, octadecyl silica; HPLC, high-performance liquid chromatography; NR, not reported;
SFE, supercritical fluid extraction
A3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
A3.1 Natural occurrence
LAS do not occur naturally.
A3.2 Anthropogenic sources
LAS are synthetic surfactants that were introduced as prime
components of almost all types of household surfactant products in
the early 1960s to replace alkylbenzene sulfonates (ABS), which were
then in widespread use. The change-over from ABS to LAS took place
gradually, starting in the United Kingdom (1960) and then spreading
to Germany (1961), the United States of America (1963), Japan (1965)
and to other European countries (Brenner, 1968; Husmann, 1968;
Waldmeyer, 1968; Tomiyama, 1972).
After use, LAS are discharged into wastewater. As the surfactant
components of the detergent products are soluble, they eventually
reach raw sewage at concentrations of 1-7 mg/litre (Rapaport et al.,
1987). Unlike ABS, which has a branched alkyl chain structure, LAS
with a linear, straight alkyl chain structure are readily
biodegradable. Their use has alleviated significant environmental
hazards such as foaming and residual surfactant in water.
A3.2.1 Production levels and processes
Annual world production of surfactants, excluding soap, in 1990
was estimated to be about 7 million tonnes (Colin A. Houston &
Associates, Inc., 1990; Richtler & Knaut, 1991). World consumption
of LAS in 1989 was about 2.43 million tonnes, 50% of which was used
in North America, western Europe, and Japan (Hewin International
Inc., 1992). Worldwide consumption of LAS in 1990 was about 2
million tonnes, with the following geographical distribution:
western Europe, 23%; North America, 19%, eastern Asia, 16%,
South-east Asia, 12%; eastern Europe, 11%; western Asia, 7%; South
America, 7%; and Africa, 5% (CEFIC, 1992). Berna et al. (1993a)
reported that, in 1990, 380 000 tonnes were used in western Europe,
180 000 tonnes in eastern Europe, 110 000 in Africa, 100 000
tonnes in western Asia, 305 000 in eastern Asia, 180 000 in
South-east Asia, 295 000 in North America, and 140 000 in Latin
America. An additional demand of 650 000 tonnes is expected by the
year 2000. The estimates for 1990 show an increase over 1987, when
LAS production in the United States, Japan, and western Europe was
about 1.4 million tonnes, on the basis of global demand for linear
alkylbenzene (Painter & Zabel, 1988), and consumption of LAS was
about 307 500 tonnes in the United States, 485 000 tonnes in western
Europe, and 145 000 tonnes in Japan (Richtler & Knaut, 1988).
LAS are complex mixtures of isomers and homologues in
proportions dictated by the starting materials and reaction
conditions. LAS are manufactured by reacting the parent
alkylbenzenes with sulfuric acid or sulfur trioxide to give the
corresponding sulfonic acid, which is then neutralized to the
desired salt. This is usually the sodium salt but ammonium, calcium,
potassium, and triethanolamine salts are also made. The reactions
are smooth and the yields nearly quantitative. Commercial LAS
contain linear alkyl chains 10-14 carbons in length, with phenyl
groups placed at various internal positions on the alkyl chain, with
the exception of 1-phenyl (Painter & Zabel, 1988).
LAS are manufactured in an enclosed process; under normal
conditions, therefore, exposure can occur only at the stage of
detergent formulation, by inhalation or dermally. Dermal exposure is
generally short and accidental, whereas exposure by inhalation can
occur continually.
The concentration of surfactants in water from washing machines
is 0.2-0.6%. LAS are estimated to represent 5-25% of the total
surfactant mixture.
In Germany in 1988, when annual consumption of LAS in the
western states was about 85 000 tonnes, daily consumption was 3.8 g
per inhabitant per day. As consumption of drinking-water was 190
litres per inhabitant per day, the average LAS concentration in
sewage was 20 mg/litre. Consumption of LAS per capita in other
countries is shown in Table 7 (Huber, 1989).
A3.2.2 Uses
LAS are the most widely used surfactants in detergent and
cleaning products, in both liquid and powder preparations and for
household and industrial use. The amount of LAS in a product depends
on several factors, including the type of application (washing-up
products, light- and heavy-duty powders and liquids) and the
formulation, but is usually 5-25%. Small amounts of LAS are used in
non-detergent applications, but these represent less than 5% of
total world consumption.
Table 7. Specific consumption of linear alkylbenzene sulfonates (LAS)
in various countries
Country Water usage LAS usage Reference
(litres per capita (g per capita
per day) per day)
Germany - 3.8 Huber (1989)
185 2.2 Wagner (1978)
United States 560 2.6a, 2.1b Rapaport et al.
(1987)
United Kingdom 208 3.5c, 2.7c Standing Technical
Committee on
Synthetic Detergents
(1978, 1989)
Spain - 5.6a, 2.6b Berna et al. (1989)
Japan 493 2.7 Ministry of Health and
Welfare (1992);
Hewin International
Inc. (1992)
a Calculated from sales
b Calculated by analysis
c Methylene blue-active substances
A4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
Section summary
The way in which LAS enter the environment varies between
countries, but the major route is via discharge from sewage
treatment works. Direct discharge of sewage to rivers, lakes, and
the sea occurs when wastewater treatment facilities are absent or
inadequate. Another route of entry of LAS into the environment is
via disposal of sewage sludge on agricultural land.
Throughout their journey into the environment, LAS are removed
by a combination of adsorption and primary or ultimate
biodegradation. LAS adsorb onto colloidal surfaces and suspended
particles, with measured adsorption coefficients of 40-5200
litres/kg, depending on the medium and structure of the LAS. LAS
undergo primary biodegradation in all environmentally relevant
compartments, such as raw sewage, sewage treatment water, surface
waters, sediments, and soils. They are readily and ultimately
mineralized under aerobic conditions in the laboratory and the
field. They tend not to be biodegraded under methanogenic conditions
or if the initial LAS concentration is so high that microbial
degradation is inhibited (> 20-30 mg/litre). Typical half-lives for
aerobic biodegradation of LAS are 1-8 days in river water, 1-2 days
in sediments, and 5-10 days in marine systems. The rate of
biodegradation depends on temperature: biodegradation is rapid
between 10 and 25°C; at lower temperatures, biodegradation kinetics
are reduced, in close association with microbial activity. During
primary sewage treatment, LAS are partially adsorbed onto and
removed with waste sludge to an extent of about 25% (range, 10-40%).
LAS are not removed during anaerobic sludge digestion but are
removed during aerobic treatment with a half-life of about 10 days.
Application of the sludge to soil generally results in 90%
degradation within three months, with a half-life of 5-30 days.
LAS are not bioconcentrated or biomagnified in aquatic
organisms. They are readily absorbed through the gills and body
surface of fish and are distributed via the blood to the systemic
organs. Most LAS-related compounds (parent compound and metabolites)
have been detected in the gall-bladder and hepatopancreas of fish.
They are usually cleared rapidly, with a half-life of two to three
days.
A4.1 Transport and distribution between media
Detergent chemicals such as LAS are normally discharged after
use into sewers in communal wastewater. The proportion of wastewater
that is subjected to sewage treatment varies widely between
countries. In most advanced countries, 50 to > 90% may be treated,
whereas in less developed countries the proportion may be as little
as 5-30% (Eurostat, 1991). In countries where there is no or
inadequate sewage treatment, LAS are removed from the environment
via adsorption and mineralization in the receiving surface waters.
Anionic surfactants such as LAS can adsorb onto the solid
substrates associated with sewage, sludge, sediment, and soil; the
extent of adsorption is dependent on the composition and physical
nature of the solid matrix. Measured values of the adsorption
constant (Kd) for LAS on a range of solid substrates were compiled
by Painter & Zabel (1989), who reported Kd values of 590-1400
litres/kg for primary sludge, 660-5200 1itres/kg for activated
sludge, and 40-360 1itres/kg for river water sediment.
A4.1.1 Wastewater treatment
Under certain conditions, up to 50% of the LAS present can be
biodegraded in sewers before entering sewage treatment (Moreno et
al., 1990). In large-volume batch biodegradation tests with
acclimatized sludge, the MBAS levels decreased to 10% of the initial
concentration within 15 days. During biodegradation, the toxicity of
the test solution decreased in parallel with the reduction in MBAS.
A relative enrichment of the shorter chain homologues was observed
by GC analysis concurrently with the decrease in MBAS levels,
indicating preferential removal of the higher homologues (Dolan &
Hendricks, 1976).
The distribution and fate of LAS have been established in the
course of mass balance studies at sewage treatment plants in Spain
(Berna et al., 1989), Italy (Cavalli et al., 1991), Switzerland
(Giger et al., 1989), Germany (De Henau et al., 1989), and the
United States (Rapaport & Eckhoff, 1990; McAvoy et al., 1993).
Efficient, well-operated activated sludge plants generally remove
most of the LAS during aerobic treatment, and the overall removal of
LAS in primary settlement and secondary aerobic treatment stages can
be < 98% (Berna et al., 1991). Smaller amounts of LAS were
removed (77 ± 15%) in less efficient, trickling filter plants
(McAvoy et al., 1993).
The main mechanism for removal of LAS during sewage treatment is
biodegradation (Berna et al., 1991), but a significant fraction (on
average, 20-30%) of the LAS entering sewage treatment plants may be
removed on primary sewage solids and do not undergo aerobic sewage
treatment (Giger et al., 1989). Instead, the sludge is digested
under anaerobic conditions, and in some countries a high proportion
may then be applied raw or digested to agricultural land as a source
of plant nutrients (Berna et al., 1991). In Germany and the United
Kingdom, 40-45% of sewage sludge is disposed of in this way (Waters
et al., 1989). Since LAS do not undergo significant anaerobic
biodegradation under methanogenic conditions, concentrations of
3-12 g/kg can be found on dried solids in sludge (see Section 5,
Table 10). Any LAS in sludge applied to agricultural soil should
then be rapidly biodegraded, since the receiving soil environment is
aerobic. In Germany and the United Kingdom a typical application of
digested sludge was estimated to add LAS at a rate of 7-16 mg/kg
soil (Waters et al., 1989).
Adsorption can account for 15-40% of the removal of LAS from raw
sewage during the primary settlement stage of treatment (Berna et
al., 1989; Giger et al., 1989). Berna et al. (1989) reported that
precipitation and adsorption were particularly important in removing
LAS from wastewater containing high concentrations of calcium and
magnesium ions.
The percentage adsorption of C10, C11, C12, and C13 LAS
onto activated sludge, Amazon clay, and various bacteria and algae
was directly related to the chain length and phenyl position: longer
homologues and more terminal phenyl isomers were adsorbed much more
readily than other forms. Adsorption of LAS at a concentration of
23 mg/litre was found to be pH-dependent, with adsorption increasing
as the pH decreased from 7 to 3 (Yoshimura et al., 1984a).
A4.1.2 Surface waters, sediments, and soils
The half-life for the removal of LAS by combined sorption and
settling < 12 km below a sewage outfall in Rapid Creek, South
Dakota, United States, was 0.25 days. The biodegradation half-life
was 1.5 days (Rapaport & Eckhoff, 1990). The partition coefficient
of LAS between natural water and sediment was reported to increase
with increasing chain length and with the position of the phenyl
nearer to the end of the chain. Adsorption was increased when either
the concentration of suspended solids or fractional organic carbon
was increased (Amano & Fukushima, 1993).
Freshwater pearl oysters are cultivated in Lake Nishinoko,
Japan, which has an area of 2.8 km2, an average depth of 1.5 m,
and a residence period of 27 days. The water of the lake was found
to contain total concentrations of MBAS of 0.01-0.02 mg/litre and
LAS concentrations of 0.005-0.018 mg/litre. The partition
coefficients of LAS (Kd) were 70 litres/kg for bottom
sediment:water and 11 litres/kg for oyster:water (Sueishi et al.,
1988). The authors concluded that when a river flows into a
semi-enclosed lagoon, the fate of the surfactants is dominated by
mass transfer between media and transformation due to degradation
rather than spatial transportation.
LAS were present in Swiss soils that had been treated with
sludge for 10 years; however, the application rates were six times
higher than normal. The reported half-lives were 5-80 days. The
authors noted that it is not entirely correct to use half-life to
describe the loss of LAS from soils, because there is competition
between biodegradation and sorption on and into soil particulates,
and LAS may persist at very low 'threshold' levels. During the
330-day study, the levels of LAS decreased from 45 mg/kg dry soil to
a residual level of 5 mg/kg (Giger et al., 1989).
A comparison of the measured concentration of LAS with detailed
records on the amount of sludge applied on 51 fields in England
indicated that loss of LAS exceeded 98% in fields that had not
recently been sprayed with sludge; losses from fields that had been
sprayed recently were calculated to be 70-99% of the estimated
cumulative load. The calculated half-lives for removal of LAS from
soil sprayed with sludge were 7-22 days. Examination of the
distribution of homologues suggested that loss of LAS is the result
of biodegradation and not leaching (Holt et al., 1989). In a study
of the disappearance of LAS from sludge-amended soils at two
locations, the average half-lives were 26 days when sludge was
applied at a rate of 1.6 kg dry sludge per m2 (giving a
concentration of LAS of 16.4 mg/kg soil) and 33 days when sludge was
applied at 5 kg wet sludge per m2 (concentration, 52.5 mg/kg soil)
(Berna et al., 1989). In another study, the half-life for LAS in
soil was more than three months; there was no evidence that they
accumulate in soil over time (Rapaport & Eckhoff, 1990).
When C13 LAS were applied to various soil (surface) types at a
concentration of 0.05 mg/kg under laboratory test conditions, the
half-lives were 1-5 days, with an average of two days. There was no
significant variation with regard to soil type. In a second
experiment, the average half-life of C12 LAS applied to subsurface
soil was 20 days (Larson et al., 1989).
After grass, radishes, and garden beans were grown for 76 days
in soil treated with 14C-LAS at a rate of approximately
1.2 g/m2, 98% of radioactive residues were recovered, with 63.6%
released to the atmosphere, 26.8% found in the soil, 6.6% in plant
biomass, and 0.9% leached out in percolated water. When potatoes
were grown on the soil for 106 days, 97.9% of the radioactivity was
recovered, and 72.3% was released to the atmosphere, 18.3% to the
soil, 5.9% in plant biomass, and 1.4% leached into percolated water
(Figge & Schoberl, 1989).
LAS in a plume of contaminated groundwater on Cape Cod, United
States, were degraded rapidly and was found only within 0.6 km of
the sewage disposal bed (Thurman et al., 1986).
Effects on the biodegradation of LAS applied at 50 mg/litre of
aqueous dispersion were studied in three Japanese soil types
inoculated with sewage. The rate of sludge application used in this
study was not typical of that found in the environment. Primary
degradation, as measured by the presence of MBAS, reached 70% within
16 days. Addition of andosol (allophane) and weathered granite
(kaolin and illite) both reduced primary degradation, and 40-50% of
the LAS was still present after 30 days, indicating that the rate of
microbial degradation of LAS adsorbed onto soils containing large
amounts of allophane and/or sesquioxides was reduced. A
montmorillonite soil did not affect the rate of degradation (Inoue
et al., 1978).
The behaviour of C10-C13 LAS and C12 LAS at concentrations
of 50 and 100 mg/litre was studied by HPLC in perfusion tests on two
types of soil, a clay loam and a sandy loam. The sandy loam, with a
lower content of humus and clay, adsorbed less LAS with a longer
lag. During the first three days of perfusion, only adsorption
occurred, 50% being adsorbed; after nine days, decomposition was
observed and only 16.6% of the LAS remained; after 15 days, the LAS
had almost completely disappeared (Abe & Seno, 1987).
LAS were applied at a rate of 5 g/m2 to three soil types:
loamy orthic luvisol under agricultural land, sandy acidic dystric
combisol under a pine forest, and combisol irrigated with
wastewater. The half-life in loamy soil was five days; 80% was
degraded after 12 days, and none was detectable after 28 days. With
45 mm of precipitation, about 8% of the LAS percolated to a depth of
10-30 cm. The LAS moved significantly more slowly than radioactively
labelled water. LAS were less mobile in the sandy soil, with a
maximal percolation depth of 5 cm after two weeks, whereas water
percolated 15 cm. The half-life in the sandy soil was 10 days, with
80% degradation after 19 days and total degradation after 28 days.
The LAS were bound to organic material in the humic litter, which
probably slowed degradation and reduced mobility. In combisol
irrigated with wastewater, the LAS were bound mainly in the upper
5 cm, with some percolation to 10-30 cm after an application of
180 mm of wastewater. The half-life was 12 days, and 80% was
degraded within 21 days; however, there was no further degradation
after 28 days, and the remaining LAS were tightly bound. Increasing
the application rate to 50 g/m2 had no effect on percolation;
however, the half-life was doubled. Samples collected during the
winter showed much slower degradation, with half-lives of 68-117
days. Percolation was also much deeper; the authors suggest this was
due to a higher rate of precipitation and extensive evaporation
(Litz et al., 1987).
In a study of the adsorption of LAS in aqueous solution onto
clay grumusol and sandy regosol soils, a linear adsorption isotherm
was obtained. The release of the surfactant was proportional to the
initial adsorption and the soil type, suggesting ready desorption.
More LAS was adsorbed by the clay soil than by the sandy soil (Acher
& Yaron, 1977).
Hydroxy aluminium and iron adsorbed LAS more readily and with a
much larger capacity than other soil constituents, such as organic
matter, silica gel, layer silicates, and calcium carbonate (Volk &
Jackson, 1968).
In a study of the adsorption of LAS applied at a concentration
of 2 mg/litre to a variety of Wisconsin (United States) soils, a
highly significant correlation was found between adsorption and
organic matter content (including the iron and aluminium
components), phosphate fixing capacity, and aluminium content. The
removal of sesquioxides reduced the adsorption of LAS to zero;
however treatment of montmorillonitic soils with H2O2 and
Na2S2O4 increased adsorption by oxidizing and removing the
organic matter, indicating that montmorillonite can adsorb LAS.
Treatment of soils with H2O2 increased adsorption because iron
and aluminium were released from organic chelates (Krishna Murti et
al., 1966).
Adsorption of LAS to microorganisms was found on the basis of
calculated adsorption isotherms to be more important than adsorption
to humic substances (Urano et al., 1984; Urano & Saito, 1985).
A4.1.3 Fate models
One model of the fate of LAS predicted the sorption coefficient
to within one order of magnitude. The sorption distribution
coefficient was consistently underpredicted, so that when the
concentrations of LAS in interstitial and overlying water were
predicted from concentrations in sediment they were overestimated.
The model thus provided conservative estimates for assessing safety
in aquatic media (Hand et al., 1990).
The reported concentrations of LAS in Rapid Creek, South Dakota,
United States, were compared with expected concentrations generated
by the quantitative water-air-sediment interaction fugacity model,
which is based on physical, chemical, reactive, and transport
properties and emission rates into rivers. In general, close
agreement was reached: in both cases, LAS had a residence time of
about two days. The authors pointed out that the results might
differ if the model were applied in situations that differed
hydrodynamically (Holysh et al., 1986).
A model to predict surface water concentrations of LAS in German
and American rivers included the following parameters: river flow
and velocity, sewage treatment plant location and type, discharge
volume, and connected population. The values obtained were in
general agreement with those measured. The authors also investigated
a septic tank discharge at a Canadian site by applying a groundwater
model, which was based on hydrogeological, biodegradation, and
sorption data. The predicted and measured concentrations were in
good agreement (Hennes & Rapaport, 1989).
A mathematical model was derived to explain a downstream
decrease in the concentration of LAS in the Lake Teganuma estuary,
Japan. The model included the adsorption coefficient, the
biodegradation rate constant, and the rate of transport (diffusive
and settling) flux of LAS between water and sediment. The model
predictions and laboratory findings were used to confirm that
biodegradation is the predominant mechanism for removal of LAS from
the estuary (Amano et al., 1991).
A model based on data from the Lake Biwa basin was devised to
predict the fate of LAS in Japanese rivers, assuming that complete
mixing occurs in any given cross-section of a river. The parameters
included the cross-sectional mean concentration of LAS, time
elapsed, flow velocity, longitudinal dispersion coefficient, decay
due to biodegradation and sedimentation, water depth, and river
width (Sueishi et al., 1988).
The measured concentrations of LAS in United States river water
under critical flow conditions were mirrored by the predictions of a
simple dilution model, which predicts chemical concentrations below
the mixing zone of wastewater treatment plants. The model is based
on three large databases, which link river flow, treatment type and
wastewater discharge volume; the output of the model is a frequency
distribution of concentrations just below the mixing zone of
treatment plant outfalls. The model predicted that 95% of the river
waters below that point would have concentrations of LAS of <
0.35 mg/litre during critical low-flow periods. The sampling sites
selected for this study were reported to have a low dilution factor
for mixing effluent with surface water, however. The predictions
therefore represent a 'worst-case scenario', since the 95 percentile
value represents critical low-flow periods, in which the lowest ever
recorded flow is used for a consecutive period of seven days within
10 years (McAvoy et al., 1993).
A4.2 Environmental transformation
A4.2.1 Biodegradation
A4.2.1.1 Aerobic degradation
Studies on aerobic biodegradation of LAS can be divided into
those of primary degradation and those of ultimate degradation.
Primary degradation of LAS occurs during the initial reactions in
the metabolic pathway, and the products are often shorter-chain
homologues. The ultimate degradation of LAS is that of the entire
molecule to its biodegradation end-products, CO2, H2O, and
NH4. These products are used in cell synthesis or, in the case of
CO2, excreted. The ultimate degradation of LAS normally requires
the action of several species of bacteria.
The degradation pathway of LAS has been described (Huddleston &
Allred, 1963; Swisher, 1963). The steps, shown in Figure 1, are:
omega-oxidation of the end of the alkyl chain, rapid ß-oxidation of
the chain, and oxidation of the ring.
Figure 1. Postulated metabolic pathway of linear alkylbenzene sulfonates
omega-oxidation
CH3(CH2)nCH(C6H4SO3H)(CH2)mCH ------------------> COOH(CH2)nCH(C6H4SO3H)CH2)mCH3
(n>m) |
|
| ß-oxidation
|
v
ß-oxidation
COOHCH(C6H4SO3H)(CH2)mCH3 <------------------ COOH(CH2)n-2CH(C6H4SO3H)(CH2)mCH3
|
| Ring
| dihydroxylation
|
v
COOHCH(C6H2SO3H)CH2CH3 ------------------> ring fission at the 1-2 position
of the ring, then desulfonation to
aliphatic products and sulphate.
From Painter (1992)
Swisher (1981) pointed out that ultimate biodegradation (at
least 80%) is achievable under the correct conditions, which
include:
(i) the presence of mixed bacterial species,
(ii) free access to new bacteria during the test,
(iii) acclimatization,
(iv) enough growth factors and food, and
(v) limitation of the LAS concentration to that found in the
environment.
Biodegradation of LAS begins at the terminus of the alkyl chain
with an omega-oxidation and is followed by successive cleavage of
C2 fragments (ß-oxidation) (Huddleston & Allred, 1963; Swisher,
1963). The resulting sulfocarboxylic acids have a chain length of
four to five carbon atoms (Schöberl, 1989). These intermediates are
further biodegraded by oxidative scission of the aromatic ring and
cleavage of the sulfonate group (Setzkorn & Huddleston, 1965;
Swisher, 1967). Catabolites of further oxidation steps are fed into
the central metabolic pathways, i.e. the Krebs cycle and glyoxylate
cycle (Schöberl, 1989).
LAS degradation begins at the longest end of the linear alkyl
chain, with omega- and ß-oxidation, and proceeds up to the
sulfophenylmono-carboxylic acids (one to two CH2 groups) (Divo &
Cardini, 1980). Under mild conditions, as in river water,
intermediates such as sulfo-phenylcarboxylic acids are often not
degraded, as the greater distance between sulfophenyl groups and the
far end of the hydrophobic group increases the speed of primary
biodegradation (Swisher, 1976). Once the terminal methyl group has
been attacked, primary biodegradation is rapid (Swisher, 1970;
Gledhill, 1975). Short-chain sulfophenylmonocarboxylic acids were
not degraded by Pseudomonas but were degraded by mixed cultures of
microorganisms (Leidner et al., 1976). The initial attack that opens
the aromatic ring is the rate limiting step for ultimate
biodegradation: once the ring is opened, degradation is rapid.
Enzymological methods were used to show that the same sequence
of steps occurs when ring degradation proceeds via the catechol
derivative. A variety of microorganisms isolated from soil, sewage,
and river water showed at least five distinct metabolic routes for
the degradation of LAS: omega- and ß-oxidation of the side-chain;
oxidation and desulfonation followed by cleavage of the aromatic
nucleus; reductive desulfonation of the ring; and metabolic
alpha-oxidation of the side-chain, followed by ß-oxidation and
desulfonation. Metabolism of alkyl chains shorter than four carbons
was initiated through the aromatic nucleus by hydrolytic or
reductive desulfonation of the ring (Cain et al., 1971). LAS may
also be cleaved by biochemical mechanisms (Schöberl, 1989).
Primary degradation
(i) Low levels of biomass
Measurement of MBAS was compared with measurement of total
organic carbon for detecting biodegradation in shake cultures. With
the MBAS method, LAS had lost 98% of their activity within five
days, whereas 34% of the total carbon had disappeared by that time,
and 70% was lost by the end of the 31-day test (Sekiguchi et al.,
1975a).
In a modification of the screening test of the Organisation for
Economic Co-operation and Development (OECD), accepted by the
European Commission, the percentage of dissolved organic carbon was
found to have decreased by more than 80% within four weeks. The
authors cautioned that the decrease in LAS may not have been due
solely to biological degradation, since 40-50% of organic carbon was
also removed from abiotic controls, suggesting that adsorption may
account for part of the removal of LAS (Canton & Slooff, 1982). When
aerobic biodegradation of 10 mg/litre LAS was followed during a
10-day incubation period at 27°C, primary degradation, measured by
the MBAS method, was complete within 8-10 days, and the theoretical
CO2 production reached 20-25% within 10 days. At a concentration
of 20 mg/litre, no degradation was observed, but this elevated
concenration may have inhibited the microbial inoculum (Itoh et al.,
1979).
The rate and degree of biodegradation of LAS are dependent on
temperature. In an unacclimatized microbial population, no more than
25% biodegradation was achieved at 5°C during a 28-day test,whereas
at 15, 25, and 35°C about 90% degradation was achieved within 7-14
days. At 45°C, the microbial population degraded 75% of the LAS
within 14 days, but this rate of degradation was not maintained,
probably because of loss of the acclimatized seed due to the high
temperature. A clearer effect of temperature was observed when the
microorganisms were acclimatized to LAS before the test. Under these
conditions, the rate of biodegradation increased steadily with
increasing temperature from 15 to 35°C (Hollis et al., 1976).
(ii) Wastewater treatment
In the OECD screening test, there was 95% loss of LAS, measured
by the MBAS method, and similar losses were measured in OECD
confirmatory test No. 1 with 20 mg/litre LAS. In the closed-bottle
test with a concentration of LAS of 2 mg/litre, there was 90-95%
analytical loss (by the MBAS method) and 60-65% loss of biochemical
oxygen demand. Coupled-unit tests with 10 mg/litre LAS and a mean
hydraulic retention time of 6 h showed 94% removal of chemical
oxygen demand (values > 73% indicate benzene ring opening) (Fischer
& Gerike, 1975). In activated sludge, 80-90% of dissolved organic
carbon and benzene rings disappeared within 6 h (Swisher, 1972). A
bacterium similar to Klebsiella pneumoniae, isolated from sewage,
degraded 93% of a concentration of LAS reported as 1% (10 g/litre),
as measured by the MBAS method (Hong et al., 1984). A direct
correlation was found between the rate of primary degradation of
1.5 mg/litre C11.7 LAS and the initial bacterial population size
(Yediler et al., 1989).
The biodegradation of C9-C13 LAS at concentrations of 25,
50, and 65 mg/litre was monitored in activated sludge at
100 mg/litre over a period of 12 days. Four methods were used: MBAS,
chemical oxygen demand, dissolved organic carbon, and ultra-violet
spectrophotometry. The results obtained with the MBAS method showed
a percentage loss of 94-97% for the three concentrations of LAS,
whereas the other methods showed losses of approx. 50% at
25 mg/litre LAS and approx. 70% at 50 and 65 mg/litre. The specific
rate of biodegradation was calculated to be 3.6 mg/g per h, on the
basis of loss of chemical oxygen demand (Pitter & Fuka, 1979).
The degradation ratio (biochemical oxygen demand:total oxygen
demand) for LAS by a synthetic sewage solution after five days was
0.81 for a concentration of 3 mg/litre and 0.14 for 10 mg/litre.
Concentrations of 30 and 100 mg/litre LAS were not degraded during
the 14-day test. Even after acclimatization to a concentration of
5 mg/litre LAS for one month, the two higher concentrations were not
degraded, probably because these levels inhibited the microbial
inoculum (Urano & Saito, 1985).
The percentage removal of biochemical oxygen demand and of LAS
were found to be significantly correlated in activated sludge and in
a trickling filter system under laboratory and field conditions,
implying that a well-functioning sewage treatment plant effectively
removes LAS (Tang, 1974).
LAS at a concentration of 150 mg/litre were inoculated into
sewage water collected from French water treatment plants. In six
out of eight experiments, primary degradation was almost complete
(90%) within seven days, but in the other two experiments only
45-55% degradation was achieved. The authors concluded that rapid
biodegradation of LAS requires the presence of a community of
several bacterial species, including Flavobacterium, Pseudomonas,
and Acinetobacter (Gard-Terech & Palla, 1986).
In an extended aeration activated sludge plant, 95-99% of LAS
was removed. Degradation of LAS and reduction of biochemical oxygen
demand were strongly correlated, in a 1:1 ratio (Knopp et al.,
1965). In long-term laboratory tests, 95-97% of LAS was removed by
activated sludge (Janicke & Hilge, 1979).
In a wastewater treatment plant where the input water had an
MBAS concentration of 6.2-9.4 mg/litre, at least 99% of the LAS
present was removed during treatment, biodegradation accounting for
85%. The relative composition of long-chain (C12-C13) homologues
adsorbed on the suspended solids was increased in comparison with
the relative incidence of short-chain (C10-C11) homologues
detected in the aqueous phases. Sulfophenylcarboxylates were
identified as intermediates of the biodegradation of LAS but were
detected only in the aqueous and not in the adsorbed phases (Cavalli
et al., 1993b).
Biodegradation of LAS in field trials with trickling filter
sewage was 86-95%, and average biochemical oxygen demand removal was
93.8%. Thus, the LAS appeared to be removed almost as rapidly as the
naturally occurring organic material. The linear correlation between
degradation and temperature (7.5-17.5°C) was highly significant.
Further degradation (94-99%) took place after additional aeration
(Mann & Reid, 1971).
MBAS degradation did not correspond to biodegradation of LAS
(20-200 mg/litre) in laboratory sludge units, because of the
presence of intermediates not accounted for by analysis of MBAS
(Janicke, 1971).
(iii) Surface waters
Primary degradation, measured by HPLC, of 5 mg/litre C11 LAS
in a static lake microcosm was complete within 18 days. The
sulfo-phenylcarboxylic acid intermediates produced were completely
degraded within 22 days (Eggert et al., 1979).
Aerobic degradation of 5 mg/litre LAS in river water, measured
by MBAS levels, was 100% after seven days at 25°C. Under
microaerophilic conditions at 25 and 35°C), no degradation took
place within 10 days (Maurer et al., 1971; Cordon et al., 1972).
In die-away tests with water from various sites on the Tama
River, Japan, primary biodegradation (measured by the MBAS method)
was complete within 7-15 days, but total organic carbon was
completely removed within an incubation time of 45 days. In a study
of LAS in seawater collected from the mouth of the Tama River,
degradation was only 50% complete within 60 days, as measured by
total organic carbon (Sekiguchi et al., 1975b). In a study to
monitor detergent-degrading bacteria from the Han River, Republic of
Korea, the lowest density was found in January and the highest in
July; the dominant group throughout the year was Pseudomonas (Bae
et al., 1982). Mixed and pure isomers of LAS were metabolized
readily (97.5%) by bacteria collected during the summer from a
sewage lagoon, but bacteria collected from under the ice during the
winter were not able to metabolize LAS (Halvorson & Ishaque, 1969).
Primary biodegradation of C10-C13 LAS was dependent on
incubation temperature in die-away tests with water from the Tama
River, Japan: primary biodegradation was complete within two days at
27°C, within six days at 15°C, and within three days at 21°C; at a
water temperature of 10°C, however, only 20% of the LAS had been
degraded within the nine-day test (Kikuchi, 1985). The optimal
temperature for the biodegradation of LAS in a river water die-away
test was found to be 25°C (Yoshimura et al., 1984b).
Degradation of 10 mg/litre LAS in a simulated river model was
found to be almost complete within 20 days, on the basis of MBAS
levels in water and sludge; however, ultra-violet spectrophotometry
showed that 40% of the LAS remained in the water and 25% in the
sludge. LAS with an alkyl chain length of C10 were degraded more
slowly than those with a chain length of C14, and LAS compounds
with sulfylphenyl groups near the terminal part of the alkyl chains
were degraded more easily than those with such groups further from
the end (Fujiwara et al., 1975).
In a study of the biodegradation of LAS (10 mg/litre) and a 1:1
LAS:ABS (10 mg/litre) mixture in canal water with an unaerated or
aerated system, LAS were rapidly degraded in the unaerated system,
by 14.9% within two days and 40.7% within seven days. Biodegradation
was more rapid in the aerated tanks, with 40.4% degraded within two
days and 74.5% after seven days. Addition of sewage to the test
system further increased the rate of degradation in the aerated
system: addition of 0.5 ml/litre sewage resulted in degradation of
78.2% after two days and 89.4% after seven days, and addition of
1.0 ml/litre sewage resulted in degradation of 89.7% after two days
and 99.8% within three days. No results were reported for the
unaerated system. The LAS-ABS mixture was degraded more slowly than
pure LAS: after two days, 12.3% was degraded without aeration, 36.4%
with aeration, 60.1% with addition of 0.5 ml/litre sewage, and 78.3%
after addition of 1.0 ml/litre sewage. The corresponding
degradations calculated after seven days were 32.5, 66.0, 80.7, and
87.3%. The authors concluded that degradation of these detergents
was increased by aerating the tank and by increasing the number of
microflora by adding sewage (Abdel-Shafy et al., 1988).
In the Lake Teganuma estuary (Japan), an average of 66% of LAS
is removed, with seasonal variability, ranging from 28% in winter to
100% in summer. Laboratory studies (based on HPLC methods) of
estuarine water showed that LAS degraded with a half-life of eight
days at 5°C and 0.2 days at 25°C. Model calculations and field
monitoring showed that biodegradation is 10 times more important in
the removal of LAS from the estuary during summer than is the
settling of solids or adsorption to bottom sediments. At lower
temperatures, biodegradation and the other removal mechanisms are of
equal importance (Amano et al., 1991).
In well water, biodegradation of all LAS homologues
(C10-C13) and isomers (maximal concentration, 2.5 mg/litre)
after an acclimatization period of one day was reported to follow
zero-order kinetics (Yakabe et al., 1992).
In seawater, primary biodegradation of 20 mg/litre LAS was 70%
after 10 days; the half-life was six to nine days (Vives-Rego et
al., 1987).
(iv) Soil
In soil degradation tests, levels of 2.5 mg/kg MBAS were reached
within 15 days of the addition of 20 mg/kg LAS (Cordon et al.,
1972). The biodegradation of LAS in soil was studied by measuring
the amounts of ferroin reagent-active substance and total organic
carbon. At 50 mg/litre LAS, total organic carbon disappeared within
50 days, whereas total ferroin reagent-active substance was
completely lost after only 10 days. Both chemical and physical
properties of the soils affected the loss of LAS: more LAS was
adsorbed onto clay loam than sandy loam, and biodegradation occurred
more readily in the clay loam (Abe, 1984). In a further study
(initial concentration not given), loss of C10-C13 and C12
LAS was complete within 15 days when measured as ferroin
reagent-active substances; however, when measured as total organic
carbon, residues remained until day 50 in the clay loam and beyond
day 60 in the sandy loam (Abe & Seno, 1987).
Ultimate degradation
A number of studies have been conducted of the biodegradation of
phenyl-radiolabelled LAS, in which 14CO2 production was measured.
(i) Screening tests
In a simple shake-flask system with LAS, CO2 evolution reached
60% or more of the theoretical value (Gledhill, 1975).
Four gram-negative bacteria synergistically mineralized
10 mg/litre 14C-LAS. After 13 days of incubation, 29% of the
14C-LAS was mineralized to 14CO2. Pure cultures were unable to
mineralize the LAS, although three of them carried out primary
biodegradation, measured by the MBAS method (Jimenez et al., 1991).
Pseudomonas, Alcaligenes, Necromonas, and Moraxella spp. isolated
from activated sludge and river water degraded the alkyl chains of
C12 LAS, while a group of unidentified Gram-negative bacteria cleaved
the benzene ring. A mixture of the two groups degraded LAS completely
(Yoshimura et al., 1984b).
(ii) Wastewater treatment
Mixed cultures of microorganisms found under natural conditions
or in sewage treatment plants can readily degrade LAS, to 95% of
MBAS and > 80% of dissolved organic carbon (Schöberl, 1989).
During a 19-day OECD screening test for the biodegradation of
14C-LAS, there was a high degree of ring mineralization, as seen
by the evolution of 55% as 14CO2. In a continuous system, 80% of
the LAS was evolved as CO2, with a mean retention time of 3 h;
2-3% remained as unaltered surfactant and 15-25% as the
sulfophenylcarboxylic acid intermediates (Steber, 1979).
Loss of MBAS (primary biodegradation) and ring cleavage were
found to be nearly complete (> 98%) during simulated waste
treatment of 14C-LAS. During simulated secondary waste treatment,
62% of alkyl and ring carbon was converted to CO2, 28-30% was
assimilated into biomass, and 8-10% remained as soluble residue. In
die-away tests, 85-100% of the substrates of LAS were converted to
CO2 within 91 days (Nielsen et al., 1980; Nielsen & Huddleston,
1981).
Continuous-flow experiments were conducted with mixed bacterial
cultures isolated from a detergent plant wastewater containing five
species of Achromobacter and two species of Acinetobacter. All
were more efficient at primary degradation than ultimate degradation
of LAS at concentrations of 20 and 50 mg/litre. One species of each
genus could effect primary degradation even after isolation (Hrsak
et al., 1982).
In a semi-continuous activated sludge method, 95% of the phenyl
ring of radiolabelled LAS was cleaved, indicating near complete
biodegradation of the whole molecule. Complete primary degradation
(MBAS method) of C10, C12, and C14 LAS was followed by 99-100%
ultimate degradation (HPLC and ultra-violet fluorescence). In
die-away tests with 10 mg/litre of C10, C12, and C14 LAS,
primary degradation was rapid and complete; 100% of C12 LAS was
removed within four days. Almost complete ultimate degradation was
observed within the 80-day test, with 90% ring cleavage of C10 LAS
and C11 LAS within 10 days and 70% ring cleavage of C14 LAS
within 30 days; however, no HPLC analysis was carried out on C14
LAS after day 30 (Huddleston & Nielsen, 1979).
The biodegradation of LAS (C9-C14) by a mixed bacterial
culture was studied in river water, forest soil, and wastewater from
a detergent plant. The bacteria were acclimatized to 10 mg/litre
LAS. Under continuous-flow conditions, LAS at a concentration of
20.8 mg/litre were 96% degraded, and a concentration of 46 mg/litre
was 64% degraded. Only 8-10% of the breakdown products were
completely mineralized; however, under the flow-through conditions
of this test, water-soluble compounds were usually removed via the
aqueous effluent and were not present long enough to allow
mineralization. Acclimatization considerably increased the kinetics
of mineralization (Hrsak et al., 1981).
(iii) Surface water and sediment
Detritus is a significant site of surfactant removal, and LAS
were found to be the most sorptive of the surfactants tested. In
wastewater from a pond containing submerged oak leaves, degradation
followed first-order kinetics, with a half-life of 12.6 days. LAS
were mineralized more slowly by leaves from a control pond, and an
S-shaped pattern of degradation was seen (Federle & Ventullo, 1990).
In river water in which the biomass levels were 10-100 times
higher below than above a sewage outfall, primary degradation of
added C11.6 LAS (11 mg/litre) and background LAS (0.37 mg/litre)
was rapid in water taken from below the outfall, with a half-life of
0.23 days (based on measurements of MBAS). Mineralization of the
benzene ring was rapid in water from below the outfall containing
sediment (500 mg/litre), with a half-life of 0.7 days. Water taken
above the sewage outfall also underwent ring mineralization, but the
rate of degradation was about 25% of that seen for water from below
the outfall, with a half-life of 2.7 days. When samples were
incubated in the absence of sediment, ring degradation was much
slower, with half-lives of 1.4 days in water taken from below the
outfall and approx.14 days in water taken above it. In all cases,
degradation was immediate in water taken below the outfall, but
occurred after a three-day lag in water taken above (Larson & Payne,
1981).
Degradation of C10-C14 homologues of LAS at concentrations
of 10 or 100 µg/litre followed first-order kinetics in both river
water and river water plus sediment; the half-time for
mineralization of the benzene ring was 15-33 h. The length of the
alkyl chain and the phenyl position had no significant effect, and
there was no effect of suspended sediment or competing homologues
(Larson, 1990).
LAS were degraded in leaf litter, creek water, periphyton, and
sediment at temperatures as low as 4°C, with half-lives of 6-11
days. Temperature changes altered the dependence of the
biodegradation of LAS: the half-lives increased by less than a
factor of two over an 18°C temperature range. Under realistic
conditions, temperature had less effect than was predicted on the
basis of classical thermodynamic studies in the laboratory
(Palmisano et al., 1991). The dependence of the biodegradation of
LAS follows a classical Arrhenius relationship down to about 12°C,
with a tenfold increase in reaction kinetics for every 2°C drop in
temperature (Larson, 1990).
Mineralization of LAS in saturated subsurface sediment from a
wastewater pond and in a pristine pond was monitored by amending the
sediment with 14C-LAS and measuring the evolution of 14CO2.
Mineralization in both sediments exhibited first-order kinetics. LAS
were mineralized without a lag in wastewater sediment, with
half-lives of 3.2-16.5 days. In the control pond, LAS were
mineralized much more slowly, with half-lives of 5.2-1540 days, and
only after a lag of 2-40 days; the lag tended to increase with
increasing depth. These findings confirm the assumption that
acclimatization considerably increases the kinetics of LAS
mineralization (Federle & Pastwa, 1988).
A study was conducted of the biodegradation of LAS by
microorganisms associated with the roots of two aquatic plants,
duckweed (Lemna minor) and cattail (Typha latifola).
Microorganisms from the roots of cattail mineralized 14C-LAS
without a lag, attaining 17% evolution of 14CO2 within the
35-day experiment. Microbiota associated with duckweed roots did not
mineralize LAS. The fact that the plants came from a pristine pond
or from a wastewater pond had no effect on the ability of the
microorganisms to mineralize LAS (Federle & Schwab, 1989).
More than 70% of parent LAS (20 mg/litre) in natural seawater at
22°C was biodegraded within 10 days, with an estimated half-life of
6-9 days (Vives-Rego et al., 1987). In an investigation of the
primary biodegradation kinetics of LAS (10 mg/litre) in natural
seawater in the presence of sediments (250 g/litre), 60% remained
after 20 days at 15°C and almost 100% of LAS at 5°C; however, at 20
and 25°C, only a small percentage of the original concentration
remained (Sales et al., 1987). In another study in seawater, 97% of
parent LAS (10 mg/litre) was biodegraded within two weeks (von Bock
& Mann, 1971).
More than 85% of LAS (C11.8) in estuarine water underwent
primary biodegradation, measured as MBAS removal, after 11 days
(Arthur D. Little Inc., 1991). In water from Chesapeake Bay, United
States, 75% of MBAS were removed within three days (Cook & Goldman,
1974). In a study of effluent-exposed estuarine waters, with
phenyl-radiolabelled C13 LAS, production of 14CO2 represented
42% of the added label. Addition of sediment from the site
(1 g/litre) increased the 14CO2 yield to 60%. In both tests, the
half-life for mineralization of LAS was about seven days. Up to 54%
of a radiolabelled control chemical, glucose, was mineralized. Thus,
mineralization of LAS occurs rapidly in pre-exposed estuarine
systems, with half-lives shorter than the typical hydraulic
residence times of such estuaries (Shimp, 1989).
(iv) Soils and groundwater
A simple shake-flask system was used to determine CO2
evolution in a test to assess the ultimate biodegradability of LAS
by microorganisms in soil and sewage. At 30 mg/litre, high
relative-molecular-mass LAS were biodegraded more slowly than those
with a low relative molecular mass. Ultimate biodegradation could
not be assessed precisely within the 28-day test period, but CO2
removal was 37-77% and dissolved organic carbon removal was 59-84%.
Ultimate biodegradation of the entire molecule (total CO2)
occurred concomitantly with biodegradation of the benzene ring
(14CO2). Ring desulfonation, measured as 35S-LAS, was rapid
and occurred mainly after primary biodegradation (MBAS method)
(Gledhill, 1975).
The kinetics of the ultimate biodegradation of C10-C14 LAS
to CO2 was studied in a sludge-amended soil at 0.1-10 times
environmental concentrations. All four homologues underwent rapid
degradation, with half-lives for the breakdown of the benzene ring
of 18-26 days (Ward & Larson, 1989).
Microbial mineralization of 50 µg/kg 14C-LAS was examined in
soil types ranging from a loamy sand impacted with sewage effluent
to a highly organic alpine soil, by monitoring the evolution of
14CO2. LAS were mineralized without a lag in all soils;
mineralization exhibited first-order kinetics in nine of the 11 soil
types. Asymptotic yields of CO2 ranged from 16 to 70%; the
half-lives were 1.1-3.7 days. The degradation rates were not
correlated with microbial activity, pH, total organic content, or
previous exposure (Knaebel et al., 1990).
After 14C calcium and sodium salts of LAS were applied to two
silty loam soils, the distribution of 14C was similar. After 60
days, 31-47% of the applied 14C had evolved as 14CO2 and
31-40% was present as soil residue, possibly as a combination of
parent and metabolized surfactant (Kawashima & Takeno, 1982).
A4.2.1.2 Anaerobic degradation
Degradation of LAS (measured as MBAS) was much slower under
anaerobic conditions in activated sludge than under aerobic
conditions. No degradation had taken place after one day; up to 20%
had been degraded between days 3 and 21, and 36% after 28 days. When
soil and wastewater were used, only 20% of the MBAS had disappeared
within 28 days (Oba et al., 1967). No significant removal of LAS was
reported in an anaerobic sludge digester at a Swiss sewage treatment
plant (Giger at al., 1989).
In a review of the fate of LAS in anaerobic and aerobic sewage
treatment plants, it was concluded that drying anaerobic sludge on
open beds considerably reduces the LAS content. Anaerobic
degradation of LAS is, however, limited, as the addition of LAS at
15 g/kg raw sewage (about 15 g/litre raw sewage) may inhibit
anaerobic degradation. In the laboratory, digestion of LAS was
impaired at concentrations of > 15-20 g/kg, and a concentration of
20 g/kg seriously inhibited gas production, especially when other
potentially inhibitory compounds were present. The concentration of
LAS normally found in sewage (5-10 g/kg) is, however, unlikely to
inhibit anaerobic degradation (Painter & Zabel, 1989). About 15-35%
of LAS in raw sewage is physically removed in primary settlers in
sewage treatment plants, accounting for most of the LAS found in
anaerobic sludge. Precipitation of LAS is correlated with water
hardness, since the solubilities of the calcium and magnesium salts
of LAS are very low; the solubility products ranged from
2.2 × 10-10 for C10 LAS to 6.2 × 10-13 for C13 LAS (Berna et
al., 1989). The effect of water hardness was confirmed by mass
balance analysis of Na+, Ca2+, and Mg2+ (Berna et al., 1993b).
The content of total calcium and magnesium in anaerobically digested
sludge was 43 times higher than that in water. High contents of LAS
in the sludge (up to 30 g/kg) did not inhibit the anaerobic
digestion process (Painter & Mobey, 1992), probably because LAS were
present as calcium and magnesium salts and therefore had reduced
bioavailability.
LAS were not degraded in an anaerobic sediment from a pond
receiving wastewater from a laundromat. Despite an exposure period
of 25 years, no anaerobic degradation was reported (Federle &
Schwab, 1992).
Pre-aerobic treatment of LAS may cause changes in the molecule
that permit subsequent degradation under anaerobic conditions (Ward,
1986).
A4.2.2 Abiotic degradation
The mechanisms of abiotic degradation of LAS reported below are
not of environmental significance, since biodegradation and sorption
are rapid, effective removal mechanisms.
A4.2.2.1 Photodegradation
In a study of the kinetics of the photodecomposition of C12
LAS, using a continuous-flow reactor, the initial concentrations
were 60-182 mg/litre and the radiation wavelength was 200-450 nm.
Conversion of LAS to intermediate products occurred within 1 min,
yielding 7 mol CO2 per mol LAS, and was complete within 20 min.
The reaction rate was increased by two orders of magnitude by ferric
perchlorate (Matsuura & Smith, 1970).
Rapid photodegradation of LAS (50 mg/litre) occurred within
1-2 h in an aqueous, aerated titanium dioxide suspension without
noble metal catalysts. There was rapid decomposition of the aromatic
ring and slower oxidation of the aliphatic ring. Photodegradation
was dependent on the simultaneous presence of titanium dioxide,
oxygen, and light (Hidaka et al., 1985).
A4.2.2.2 Cobalt-60 irradiation
The decomposition of LAS was studied in distilled water
irradiated with cobalt-60 gamma rays, which react with water to
produce oxygen, peroxide, hydrogen peroxide, and other strong
oxidizing agents. A concentration of 10 mg/litre LAS was reduced to
7.8 mg/litre by absorption of 10 Gy and to 0.9 mg/litre by
absorption of 100 Gy. The rate of irradiation was found to be less
important than the total absorbed energy (Rohrer & Woodbridge,
1975).
A4.2.3 Bioaccumulation and biomagnification
Studies of the bioaccumulation potential of LAS have all been
carried out with LAS labelled with 14C or 35S. It should be
noted that as these techniques do not usually allow consideration of
metabolic transformation the actual bioaccumulation of the parent
compound may be overestimated. Toxic concentrations of the breakdown
products of LAS are discussed in section A9.3.7.
A4.2.3.1 Aquatic organisms
Bioaccumulation has been studied in daphnids and fish (Table 8).
LAS are readily absorbed through the gills and body surface of fish
and are subsequently distributed via the blood to the organs and
tissues; most LAS accumulate in the gall-bladder and hepatopancreas.
Clearance is usually rapid, with a half-life of two to three days.
Short-chain LAS are accumulated to a lesser degree than long-chain
LAS.
Only 1% of 0.5 mg/litre LAS added to water was retained in
Daphnia magna within three or four days after transfer to 'clean'
water. Almost all of the chemical was in the form of intact LAS. In
fathead minnows (Pimephales promelas), metabolic transformation
occurred. All tissues monitored showed some uptake, with
concentration factors ranging from 79-372 in muscle to 21 000-70 000
in gall-bladder. Within four days of transfer to 'clean' water, 85%
of the LAS had been lost, and almost 100% was lost within 10 days
(Comotto et al., 1979).
Table 8. Bioconcentration factors for linear alkylbenzene sulfonates in aquatic invertebrates and fish
Organism Static/flow Exposure Duration Chain Steady Bioconcentration Tissue Reference
concentration of test length state factor
(mg/litre) (days)
Daphnia magna Flow 0.07 3 C12 ? 490 Comotto et al.
(1979)
0.11 560
044 720
0.09 3 C13 Yes 1250
0.11 1050
0.41 1325
Cyprinus carpio Static 61.1 1 C12 Yes 4.1 Skin surface Kikuchi et al.
(1978)
1000 Gall-bladder
Flow 0.5 4 C12 Yes 20 Whole body Wakabayashi
30 Hepatopancreas et al. (1978)
9000 Gall-bladder
0.0091 5 C12 Yes 16 Whole body Wakabayashi
et al. (1981)
0.3 400
1.0 300
Pimephales Flow 0.1 11 C12 Yes 551 Whole body Comotto et al.
promelas C13 1223 (1979)
C12, C13 269
Table 8 (contd)
Organism Static/flow Exposure Duration Chain Steady Bioconcentration Tissue Reference
concentration of test length state factor
(mg/litre) (days)
Lepomis Flow 0.063 28 C12 Yes 260 Whole body Bishop &
macrochinus 0.064 120 Maki (1980)
Flow 0.5 35 C11.7 Yes 107 Whole body Kimerle et al.
5000 Gall-bladder (1981)
Static, water unchanged for the duration of the test; flow, concentration in water maintained continuously
In an experiment in which the aqueous concentrations of an
initial concentration of 1.1 mg/litre LAS decreased by 20% during
the test, the compounds were concentrated in the gills of carp
(Cyprinus carpio) within 2 h of exposure, with a concentration
factor of 40. Skin surface, muscle, brain, kidney, hepatopancreas,
and gall-bladder showed more gradual uptake of LAS over the 24 h of
exposure, with concentration factors ranging from 4.1 for skin
surface to 1000 for gall-bladder. Blood, gonads, and spleen also
took up LAS but were not monitored throughout the period of
exposure. LAS was lost rapidly from all tissues except the
gall-bladder during 48 h in 'clean' water (Kikuchi et al., 1978).
In the bluegill (Lepomis macrochirus), a steady state was
reached within 120-168 h. The bioconcentration factor was calculated
by a kinetic method to be 286 for a concentration of LAS of
0.8 mg/litre and 132 for 0.08 mg/litre. LAS were cleared rapidly
after the fish were transferred to 'clean' water, with 99%
eliminated within 336 h; the time for clearance was 29-30 h (Bishop
& Maki, 1980). In another study, a steady state was reached within
seven days; the bioconcentration factor in whole body using a
kinetic method was reported to be 104; and the half-time for
clearance was two to five days during a depuration period of 14
days. The authors postulated that fish excrete LAS in the urine and
excrete shorter-chain carboxylates with the benzene ring intact
across the gill membranes. Both forms may also be excreted in the
faeces (Kimerle et al., 1981).
A4.2.3.2 Terrestrial plants
Foliar uptake of the calcium and sodium salts of 14C-LAS
(chain length not specified) by peanuts was studied seven and 30
days after application. No movement of LAS was detected: 70-80%
remained within the same leaf to which the compound was applied, and
no LAS were detected in other parts of the plant (Kawashima &
Takeno, 1982).
Aqueous solutions of 14C-LAS (chain length not specified) were
applied to soil (orthic luvisol), and ryegrass (Lolium perenne)
was grown under laboratory conditions for up to seven days. Uptake
of LAS after three days was 80 mg/kg at an application rate of
1 mg/kg dry weight, 370 mg/kg at a rate of 5 mg/kg, and 18 900 mg/kg
at 50 mg/kg. After seven days, levels of 600, 5000, and 19 300 mg/kg
were measured at the three dose levels, respectively (Litz et al.,
1987).
14C-LAS (chain length not specified) were applied under field
conditions to both loamy orthic luvisol and sandy dystric cambisol
soils irrigated with wastewater at rates of 5 and 50 g/m2. After
49 days, rye grass grown in the loamy soil contained residues of 130
and 1000 mg/kg dry weight at the two exposure rates, respectively.
Plants grown in the sandy soil contained 230 and 470 mg/kg,
respectively, after 54 days (Litz et al., 1987).
Two plant-soil microcosms were exposed to 14C-LAS (chain
length not specified), and LAS degradation and percolation were
followed for up to 109 days. The initial soil concentrations were
16.2 µg/g dry soil in potato soil (sandy) and 27.2 µg/g in grass,
bean, and radish soil (clay-like). The concentrations of
radiolabelled compounds in the plants decreased rapidly: at the end
of exposure, 39.1-65.8 µg LAS equivalents per g fresh weight of
plant were found in potatoes (study duration, 76 days) and
62.1-213.3 µg/g in grass, radishes, and beans (study duration, 109
days) (Figge & Schoberl, 1989).
A5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
Section summary
The concentrations of LAS have been quantified by means of a
specific, sensitive analytical method in almost every environmental
compartment in which they might be present. The concentrations
decrease progressively from wastewater to treated effluent and
surface waters, and low concentrations are found in the sea.
The environmental concentrations of LAS are directly dependent
on use patterns, the type and efficiency of sewage treatment, and
the characteristics of the receiving environment. In areas where LAS
are the predominant surfactants used, typical concentrations are
1-10 mg/litre in wastewater, 0.05-0.1 mg/litre in effluents that
have undergone biological treatment, 0.05-0.6 mg/litre in effluents
passed through a percolating filter, 0.005-0.050 mg/litre in surface
waters below sewage outfalls (with concentrations decreasing rapidly
to 0.01 mg/litre downstream from the outfall), < 1-10 mg/kg in
river sediments (up to 100 mg/kg in highly polluted sediments near
discharge zones), 1-10 g/kg in sewage sludge, and < 1-5 mg/kg in
sludge-amended soils. The initial concentration of LAS in
sludge-amended soils is 5-10 mg/kg, but up to 50 mg/kg have been
reported after atypically heavy appli-cations. The concentration of
LAS in estuarine waters is 0.001-0.010 mg/litre but is higher where
wastewater is discharged directly. The concentrations in offshore
marine waters are < 0.001-0.002 mg/litre.
A wide range of environmental concentrations has been reported,
owing to use of different analytical methods; differences in
characteristics of sampling sites, which range from highly polluted
areas with inadequate sewage treatment to areas where sewage
undergoes extensive treatment; seasonal differences, which can
account for a twofold variation; and differences in the use of LAS.
Monitoring has shown no accumulation of LAS in environmental
compartments over time. The concentrations in soil do not increase
with time but are diminished due to mineralization. As LAS are not
degraded under strictly anaerobic condition, they are not
mineralized in anaerobic sediments. With current use of LAS, the
rates of their assimilation in all receiving environmental
compartments is equal to their rate of input, implying a steady
state.
A5.1 Environmental levels
LAS have been measured in most environmental compartments,
including discharge (raw sewage), sewers, sewage treatment plants,
sludge-amended soils and land fill, river water, river sediments,
subsurface soils, groundwater, and estuaries (Berna et al., 1991).
A decline in the concentrations of anionic surfactants in the
environment, as assessed by measurement of MBAS, was seen in Europe,
Japan, and the United States after ABS was replaced by LAS (Sullivan
& Evans, 1968; Sullivan & Swisher, 1969; Gerike et al., 1989).
Similar declines have been observed more recently in countries such
as Thailand, where the change to LAS detergents is also more recent
(Berna et al., 1991).
A5.1.1 Wastewater, sewage effluent, and sludge
The concentrations of LAS in sewage influent and effluent at
sewage treatment plants are shown in Table 9; those in sewage sludge
are given in Table 10.
The efficiency of wastewater treatment plants in removing LAS is
reported to exceed that of removal of biochemical oxygen demand.
Activated sludge removed an average of 98% LAS, trickling filters
removed 80%, and primary clarification, 27%. The average
concentration in raw sewage was 3.5 mg/litre, and those in effluent
were 2.1 mg/litre after primary treatment and 0.06 mg/litre in
activated sludge. The average chain length of LAS was C12.5 in
sewage sludge and C12 in influent sewage (Rapaport & Eckhoff,
1990).
The amount of LAS removed in a sewage treatment plant was 93% on
the basis of total organic carbon and 98.1% on the basis of a
specific method. The contribution of LAS to the total organic carbon
was estimated to be 0.93% in treated water and 3.0% in digested
sludge; 75.9% of LAS present in the raw sewage was mineralized
during treatment and 7% was in the form of sulfoxyphenyl-
carboxylates, a product of the biodegradation of LAS, suggesting
that biodegradation of LAS had reached a steady state. These figures
were obtained by analysis for sulfoxyphenylcarboxylates (Berna et
al., 1993b).
In another study, 40% of LAS was removed in a wastewater
treatment plant. The half-life for removal from the sewer pipe was
calculated to be 11 h (Moreno et al., 1990).
A5.1.2 Sediment
The concentrations of LAS in sediment are shown in Table 11, and
those in sediment samples collected at various distances from sites
of effluent outfall are shown in Table 12.
Table 9. Concentrations of methylene blue-active substances (MBAS) and linear alkylbenzene sulfonates (LAS) in sewage
influent and effluent
Location Year Material Concentration (mg/litre) Reference
MBAS LAS
Switzerland (29 sites, 1986 Raw sewage 0.95-3.9 Brunner et al. (1988)
1 sampling) Effluent 0.007-0.33
Germany (11 sites, 1985 Influent (activated sludge) 5.1 (1-13.3) 4 (0.54-12.4) Matthijs & De Henau
1 sampling) Influent (trickling filter) 8.8 (8.1-9.9) 7.4 (6.8-8.4) (1987)
Effluent (activated sludge) 0.19 (0.09-0.28) 0.07 (0.05-0.11)
Effluent (trickling filter) 1.1 (0.84-1.5) 0.76 (0.61-0.94)
United Kingdom
(several samples) 1982 Effluent 0.69 (0.58-0.81) 0.31 (0.21-0.42) Gilbert & Pettigrew
(1984)
River Thames area 1987 Sludge 15.1-341 Holt et al. (1989)
(5 sites, several
samples)
Israel (4 sites) 1983 Influent 9.6-10.6a Zoller (1985)
Effluent 0.3-11.0a
United States 1979 Effluent 0.078-0.303 Eganhouse et al. (1983)
(4 sites, 45 samples 1976-86 Influent 3.7 ± 1.1 Rapaport & Eckhoff
Effluent (activated sludge) 0.05 ± 0.04 (1990)
Effluent (trickling filter) 0.6 ± 0.3
Effluent (primary) 2.2 ± 0.4
Table 9 (contd)
Location Year Material Concentration (mg/litre) Reference
MBAS LAS
United States Influent 5.9-6.5 5.7-6.5 Osburn (1986)
(1 sampling) Influent 3.7-5.2 3.8-4.9
Effluent 0.39-1.02 0.14-0.60
(2 sites, 9 samples) 1983 Raw influent 4.17 3.73 Sedlak & Booman
Primary influent 3.18 2.97 (1986)
Primary effluent 1.66-2.82 1.73-2.51
Final effluent 0.03-0.06 0.02-0.05
Canada (4 sites, 1976-86 Influent 2.0 ± 0.6 Rapaport & Eckhoff
45 samples yearly) Effluent (activated sludge) 0.09 ± 0.05 (1990)
Effluent (primary) 1.7-2.3
Japan
(5 sites, 60 samples) 1972-73 Influent 5.1-14.0 Oba et al. (1976)
Effluent 0.3-4.7
(6 sites, 1-2 samples) 1984 Influent (suspended particles) 0.236-1.504 Takada & Ishiwatari
Effluent (suspended particles) 0.0001-0.001 (1987)
a Total anionic surfactants (mainly LAS)
Table 10. Concentrations of methylene blue-active substances (MBAS) and linear alkylbenzene sulfonates (LAS) in sewage sludge
Location Year Material Concentration (mg/litre) Reference
MBAS LAS
Switzerland
(8 and 12 sites, Digested sludge 2900-11 900 McEvoy & Giger
(1985, 1986)
1 sampling)
(29 sites, 1986 50-13 800a Brunner et al.
1 sampling) (1988)
Spain
(5 sites, several Activated sludge 7000-30 200a Berna et al. (1989)
samplings) (anaerobic digestion)
Aerated, settling system 400-700a
Finland (12 sites, Digested sludge 3400-6300a McEvoy & Giger
1 sampling) (1986)
Belgium (11 sites, 1985 Aerobic sludge 5399 (3042-8133) 281 (182-432) Matthiijs & De
1 sampling) Digested sludge 9017 (3632-17 006) 4917 (1327-9927) Henau (1987)
Germany (4 sites, 1981-86 4920 (1330-9930) Rapaport &
45 samples yearly) Eckhoff (1990)
Table 10 (contd.)
Location
Year Material Concentration (mg/litre) Reference
MBAS LAS
United States
(4 sites, 45 1981-86 4660 ± 1540 Rapaport &
samples yearly) Eckhoff (1990)
12 sites, NY, Digested sludge 6900a McEvoy & Giger
(1 sampling) (1986)
(12 sites, CA, Digested sludge 5200a
1 sampling)
(1 sampling) Primary sludge 110-126 107-127 Osburn (1986)
(2 sites, OK, 1983 Primary sludge 4610-6120 5340-6310 Sedlak & Booman
9 samples) Secondary sludge 520-990 410-860 (1986)
Anaerobic digester 6860 6660
Aerobic digester 3820 4250
Drying bed (anaerobic) 170 160
Drying bed (aerobic) 230 150
Southern California 1981 Effluent particulates 1342 Eganhouse et al.
(marine) (1983)
a Dry weight
Table 11. Concentrations of methylene blue-active substances (MBAS) and linear alkylbenzene sulfonates (LAS) in sediments in
the United States and Japan
Location Year Concentration (mg/kg) Reference
MBAS LAS
United States
Rivers (activated sludge) 0.3-3.8 McAvoy et al. (1993)
Rivers (trickling filter) 0.2-340
Mississippi River 1991-92 < 0.01-5 Tabor et al. (1993)
Japan
Tokyo Bay (1 sampling, few samples) 1969 35 (33-37) Ambe (1973)
River (1 sampling, few samples) 61 (55-65)
River Sagami estuary (16 sites, 1 sampling) 7.9-39a ND-17 Utsunomiya et al. (1980)
Sagami Bay (16 sites, 1 sampling) 5.1-15 ND
Rivers 1977 < 1-260 Environment Agency Japan
(1978)
Lake Suwa (1 site, 3 samples) 1977 1.0-7.0
Rivers (9 sites, 7 samples, 1 year); 1982-83 107 (ND-567) Takada & Ishiwatari (1987);
(1 site 52 samples) Takada et al. (1992b)
Estuaries (1 site, 52 samples) 1983-84 4.82 (0.12-36.6) Takada et al. (1992b)
Tokyo Bay (9 sites, 7 samples, 1 year) 1980 71.0 Takada & Ishiwatari (1987)
Tokyo Bay 1984 0.02 (ND-0.06) Takada et al. (1992a)
Sumida River (12 sites, 1 sampling) 1982 0.069 Kikuchi et al. (1986)
Tama River (3 sites, 8 samples) 1977 3.5-86.3 Hon-Nami & Hanya (1980b)
Tama River (10-12 sites) 1982 0.141 Kikuchi et al. (1986)
Tokyo Bay (10-12 sites) 1982 < 0.001-0.002
Table 11 (contd)
Location Year Concentration (mg/kg) Reference
MBAS LAS
Japan (contd).
Tsurumi River (7 sites, 12 samples) 1984 17-45a Yoshikawa et al. (1985)
Tama River 1981 2.79-10.72 Yoshimura et al. (1984b)
Ports and coast 1977 < 1-2.9 Environment Agency Japan
(1978)
ND, not determined
a Dry weight
Table 12. Concentrations of methylene blue-active substances (MBAS) and linear alkylbenzene sulfonates (LAS) in sediment of rivers in
Germany and the United States at various distances from effluent outfalls
Location Year Sampling site (distance Concentration (mg/litre) Reference
from effluent outfall)
MBAS LAS
German rivers (14 sites, several 1978-82 Below outfall 1.5-174a De Henau et al.
samples) (1986)
United States
Rivers (4 sites, 45 samples) 1978-82 Below outfall 190 Rapaport &
yearly < 5 miles (8.0 km) 11.9 Eckhoff
> 5 miles (8.0 km) 5.3 (1990)
(1 sampling) 0.5 miles (0.8 km) 118-317 100-322 Osburn (1986)
4.4 miles (7.1 km) 4.1-19 2.0-5.1
7.4 miles (11.9 km) 7.5-10.6 1.3-4.4
Rapid Creek, South Dakota 1979-80 0.8 km 44.6-275 Games (1983)
7 km 3.2-9.1
11.7 km 2.1-8.4
25.3 km 2.7-10.1
48 km 1.4
87.2 km 1.5
Little Miami River, Ohio Downstream from sewage ND-1.2 Hand et al. (1990)
4 sites, 1 sampling) treatment plants 24.7-290b
Rivers (4 sites, 45 samples) 1978-82 Below outfall 190 Rapaport &
Above outfallc 1.0-1.2 Eckhoff (1990);
Below outfall (left)c 0.3-1.6 McAvoy et al.
Below outfall (middle)c 0.6-3.8 (1993)
Table 12 (contd)
Location Year Sampling site (distance Concentration (mg/litre) Reference
from effluent outfall)
MBAS LAS
United States (contd)
Rivers (4 sites, 45 samples) 1978-82 Below outfall (right)c 0.8-3.4
(contd). Above outfalld 0.2-0.9
Below outfall (left)d 0.2-130
Below outfall (middle)d 0.6-124
Below outfall (right)d 9-340
a 13 of the 14 samples contained < 25 mg/kg and 10 contained < 10 mg/kg
b Suspended solids
c Activated sludge
d Trickling filter
Concentrations of LAS > 10 mg/kg were measured in sediments
from the upper estuaries near Tokyo Bay and < 1 mg/kg in the lower
estuaries. The concentrations of LAS in sediments decreased
offshore, falling below 0.01 mg/kg in sediments sampled 10 km from
the mouths of the rivers. The authors suggested that loss of LAS was
due to rapid degradation in the coastal zone (Takada et al., 1992a).
It was reported in one study that C13 was the most abundant
homologue of LAS in river sediment (Yoshikawa et al., 1985); another
group found that C12 was the most abundant of the LAS in estuarine
sediments and that no C10 were present (Utsunomiya et al., 1980).
C12 and C13 LAS predominated in sediment and C10 and C11
homologues were the most abundant in water (Hon-Nami & Hanya,
1980b). The average chain length of LAS in Japanese river sediments
was C11.8-C12.2 (Hon-Nami & Hanya, 1980b; Yoshimura et al.,
1984a).
In a study of marine sediments from an area adjacent to the
point of discharge from a submarine sewer, LAS were detected only in
the vicinity of the discharge, at a concentration of 0.1 mg/kg, and
not in sediment sampled 50 m outside this area. The average chain
length was C11.7. In a comparison of the chain lengths of LAS
detected in various environmental compartments and those used in
detergent products, the LAS detected in sludge and sediment were
relatively higher homologues and those in the water phase were
lighter (Prats et al., 1993).
The average concentration of LAS in river sediments sampled
upstream of an activated sludge treatment plant outfall was
1.1 mg/kg, and those in sediments downstream of the plant were
0.3-3.8 mg/kg (McAvoy et al., 1993).
A5.1.3 Surface water
The concentrations of LAS in water are shown in Table 13 and
those in samples taken at various distances from sites of effluent
outfall in Table 14.
After replacement of branched-chain ABS, which are only
sparingly biodegradable, with the straight-chain LAS, the
concentrations of MBAS decreased in many rivers. ABS were replaced
by LAS in Japan in the late 1960s; the ratio of LAS to total ABS in
river water rose from 20 to 70% in 1967-70 and had reached 90% by
1973 (Miura et al., 1968; Ihara et al., 1970; Oba et al., 1975). The
levels of MBAS were monitored in the Illinois River, United States,
from 1959 to 1966; those in 1965 and 1966 reflected the change in
surfactant usage (Sullivan & Evans, 1968), and this trend continued
in 1967 and 1968 (Sullivan & Swisher, 1969). In the River Rhine, the
level of anionic detergents, measured as MBAS, fell steadily between
1971 and 1977 (Hellmann, 1978). In water samples from 140 sites on
Table 13. Concentrations of methylene blue-active substances (MBAS) and linear alkylbenzene sulfonates (LAS) in water
Location Year Water Concentration (mg/litre) Reference
sample
MBAS LAS
Freshwater
United States
Rivers (4 sites, 45 samples yearly) 1978-86 0.041-0.115 Rapaport & Eckhoff
(1990)
Little Miami River, Ohio (4 sites, < 0.05 Hand et al. (1990)
one sampling) Interstitial ND-0.08
Illinois River (one sampling)a 1959-65 0.54 Sullivan & Swisher
1965-66 0.22 (1969)
1968 0.05-0.06
Rapid Creek, South Dakota 1979-80 0.01-0.270 Games (1983)
Mississippi River (36 sites) 1991-92 < 0.01-0.3 McAvoy et al. (1993)
(350 samples) < 0.01-0.046 < 0.005 Tabor et al. (1993)
Japan
Rivers (23 sites, 51 samples) 1977 < 0.01-2.9 Environment
Agency Japan (1978)
Rivers (1 sampling) 0.018-0.59 Tsukioka &
Murakami (1983)
Oohori River (6 sites monthly) 1987-88 approx. 0.5-1.6 Amano et al. (1991)
Lake Teganuma (6 sites monthly) 1987-88 ND-approx. 0.7
Tama River (3 sites, 8 samples) 1977-78 0.24-1.24 0.108-0.491 Hon-Nami & Hanya
(1980a)
Table 13 (contd.)
Location Year Water Concentration (mg/litre) Reference
sample
MBAS LAS
Japan (contd)
Rivers, Hyogo Prefecture (70 sites) 0.004-2.5 Kobuke (1985)
Tama River (3 sites, 1 sampling) 0.035-0.219 Yoshikawa et al.
(1984)
Tama River (10-12 sites) 1982 0.128 Kikuchi et al. (1986)
Sumida River (10-12 sites) 1982 0.005-0.01 Kikuchi et al. (1986)
Rivers (1 sampling) 0.06-0.12 Saito & Hagiwara
(1982)
Rivers, Niigata Prefecture (6 sites, 0.02-2.63 0.18 (max) Motoyama & Mukai
1 sampling)
(1981)
Rivers, coastal area, Hiroshima 0.019 Okamoto & Shirane
Prefecture (20 sites) (0.001-0.06) (1982)
Inland Sea, Eastern Seto (4 sites, 1975 0.016-0.077 Yoshida & Takeshita
1 sampling) (1978)
(17 sites, 1 sampling) 1976 0.01-0.048
Tsurumi River, Kanagawa 1984-76 Surface 0-0.8 0.01-0.29 Yoshikawa et al.
(7 sites once) (1985)
Yodo River, Osaka (several sites) 1989 Surface 0.043-0.089 Nonaka et al. (1990)
Tama River, Tokyo (2 sites, 1981 Surface 0.2 Yoshimura et al.
4 samples) (1984b)
Sumidogawa River (2 samples) 1983 Suspended 0.0048-0.054 Takada & Ishiwatari
Tomogawa River (5 samples) particles 0.0005-0.0025 (1987)
Teshiro River, Nagoya (4 sites, 1989 Surface 0.01-0.27 Kojima (1989)
4 samples)
Table 13 (contd)
Location Year Water Concentration (mg/litre) Reference
sample
MBAS LAS
Japan (contd)
Lake Biwa, Shiga 1988 Surface 0.00 Shiga Prefecture
(1988)
Teganuma, Chiba (1 site, 1988 Surface ND-0.423 Amano et al. (1989)
12 samples)
River (several sites) 1988 Surface 0.019-1.4 Nonaka et al. (1989)
Nagoya Bay 1989 Surface 0.00 Kojima (1989)
Rivers, Fukuoka City ND-1.6 Ohkuma (1981)
Europe
River Rhine (several sites) 1971-72 0.08-0.24 Hellmann (1978)
Saar River (11 sites, 1 sampling) 1985 0.13 0.04 Matthijs & De Henau
(0.03-0.25) (0.01-0.09) (1987)
German rivers (several sites) 1976-79 0.075-0.5 Fischer (1980)
Dutch river (Amsterdam drinking- 0.004-0.141 0.003-0.037 Waters (1976)
water supply) (8 sites)
Florence, Italy (several samples) 1983 Aqueduct 0 .01-0.1 Mancini et al. (1984)
(several sites) 1982 Well water 0.00-0.01
United Kigdom
Rivers 1982 0.04-0.26 0.012-0.08 Gilbert & Pettigrew
(1984)
Rivers (8 sites) 0.035-0.217 0.009-0.097 Waters (1976)
Rivers (4 sites) 1977-78 0.022-0.473 0.007-0.173 Waters & Garrigan
(1983)
Table 13 (contd)
Location Year Water Concentration (mg/litre) Reference
sample
MBAS LAS
Groundwater 1992 < 0.01-0.02 Field et al. (1992)
Estuarine and marine water
North Sea (19 sites) 1989 < 0.0005-0.0012 Stalmans et al.
(1991)
Krka River estuary, Croatia 1990 Wastewater 0.42-0.78 Terzic & Ahel (1993)
(below 50 m > 50 m) Estuarine water 0.003-0.007
0.001-0.002
Tokyo Bay, Japan (8 samples) 1978 0.03-0.07 < 0.003-0.014 Hon-Nami & Hanya
(1980a)
Tokyo Bay, Japan (10-12 samples) 1982 0.001-0.03 Kikuchi et al. (1986)
Osaka Bay, Japan (several sites) 1988 Surface ND-0.0072 Nonaka et al. (1989)
ND, not detected
a 10-20% of MBAS were LAS
Table 14. Concentrations of methylene blue-active substances (MBAS) and linear alkylbenzene sulfonates (LAS) in water at various
distances from effluent outfalls
Location Year Sampling site (distance Concentration (mg/litre) Reference
from effluent outfall)
MBAS LAS
United States
Rivers (4 sites, 45 samples 1978-86 Below outfall 0.115 Rapaport & Eckhoff
yearly) < 5 miles (8 km) 0.079 (1990)
> 5 miles (8 km) 0.041
(1 sampling) 0.5 miles (0.8 km) 0.400 0.270 Osburn (1986)
4.4 miles (7.1 km) 0.300 0.150
7.4 miles (11.9 km) 0.250 0.120
15.8 miles (25.4 km) 0.240 0.100
30.0 miles (48.3 km) 0.130 0.040
55.0 miles (88.5 km) 0.100 0.010
Rapid Creek, South Dakota 1979-80 0.8 km 0.270 Games (1983)
7 km 0.150-0.190
11.7 km 0.120
25.3 km 0.080
48 km 0.040
87.2 km 0.010
Rivers Above outfall < 0.01-0.9 McAvoy et al. (1993)
Below outfall (left) < 0.01-0.33
Below outfall (middle) < 0.01-0.3
Below outfall (right) < 0.01-0.3
Table 14 (contd)
Location Year Sampling site (distance Concentration (mg/litre) Reference
from effluent outfall)
MBAS LAS
Canadian rivers (4 sites, 45 1978-86 Below outfall 0.053 Rapaport & Eckhoff
samples yearly) (1990)
Rio Grande, Brazil 1979 90 m 0.05-4.5 Kantin et al. (1981)
(1 sampling, 50 samples)
German rivers (several sites) 1976-79 Unpolluted 0.075 Fischer (1980)
Polluted 0.2-0.5
(4 sites, 45 samples yearly) 1978-86 Below outfall 0.01-0.09 Rapaport & Eckhoff
(1990)
United Kingdom
Rivers (several samples) 1982 Above discharge 0.04 0.012 Gilbert & Pettigrew
(0.02-0.07) (0.008-0.019) (1984)
Close to discharge 0.26 0.08
(0.11-0.47) (0.01-0.17)
5-16 km 0.16 0.04
(0.08-0.23) (0.008-0.095)
Avon River (4 sites) 1977-78 Head water 0.03-0.039 0.009-0.015 Waters & Garrigan
0.5 km 0.21-0.371 0.056-0.173 (1983)
6 km 0.095-0.22 0.011-0.095
Tean River (4 sites) 1977-78 Head water 0.035-0.073 0.008-0.019
Directly below sewage 0.208-0.473 0.067-0.144
treatment
5 km 0.145-0.234 0.019-0.07
Table 14 (contd.)
Location Year Sampling site (distance Concentration (mg/litre) Reference
from effluent outfall)
MBAS LAS
United Kingdom (contd)
Trent River (4 sites) 1977-78 Head water 0.022-0.052 0.01-0.011
20-35 km below head 0.08-0.227 0.007-0.072
water
Nene River tributary 1978 0.104 0.011
(4 sites)
In the vicinity of 0.206-0.216 0.035-0.037
sewage effluent disharge
3.5 km 0.184 0.035
13.5 km 0.06 0.007
four German rivers, MBAS concentrations fell by 90% between 1964 and
1987 (Gerike et al., 1989).
The mean level of MBAS in rivers in the United Kingdom was
0.15 mg/litre. On average, only 26% was attributable to LAS (by
microdesulfonation and gas-liquid chromatography), but the levels of
LAS and their contribution to the total MBAS concentration varied
according to the sampling site, with a higher proportion of LAS in
samples from sites near sewage effluent discharge points (Waters &
Garrigan, 1983). Similar findings were reported by Gilbert &
Pettigrew (1984), who found that LAS represented 45% of total MBAS
in actual sewage. Sites immediately below sewage outfalls were found
to have higher MBAS:LAS ratios than sites further downstream
(Osburn, 1986).
In Lake Biwa basin, Japan, during the summer months of 1983, LAS
were found in a wide range of concentrations. The highest, measured
as MBAS, were > 0.2 mg/litre at river mouths. The levels in rivers
flowing from densely populated areas were 0.05-0.2 mg/litre MBAS and
those flowing from less populated areas were < 0.05 mg/litre. The
middle stream zone of the River Isasa, in a densely populated area,
contained levels of 0.36-1.91 mg/litre, and surfactant levels in
residential areas showed daily fluctuations related to discharge
(Sueishi et al., 1988). Several observations apply to these studies.
Firstly, the fact that daily fluctuations were observed indicates
that the samples may have been taken from the actual discharge
plume, so that the wastewater effluent may not have been completely
mixed with the recipient surface water. Secondly, in several
Japanese studies of heavy discharge zones, anionic surfactants could
not be detected in surface waters, although the analytical detection
limit of MBAS in the mid-1980s was 0.05-0.1 mg/litre. Thirdly,
sewage treatment at several of the sites has improved considerably
over the last decade.
Seasonal trends in the concentrations of LAS were observed in
the Oohori River and Lake Teganuma, Japan, in 1987 and 1988, with
low levels in summer and high levels in winter (Amano et al., 1991).
The concentrations of LAS were measured in the Tamagawa River,
Japan, at two-week intervals for two years, by sampling water from
the boundary between freshwater and brackish zones. The
concentrations measured in winter were about five times higher than
those measured in summer, when long-chain homologues tended to be
depleted. The distribution of isomers also showed a clear seasonal
trend, with a greater loss of external isomers in summer. The
seasonal changes are thought to be the result of differences in
water temperature and microbial activity. The flux of LAS in the
river was estimated to be 320 tons/year (293 tonnes/year), which
exceeds the total amount of LAS accumulated in the bay sediment,
indicating that > 99.9% of LAS in the estuary and the bay was
degraded (Takada et al., 1992b).
The concentrations of LAS in suspended particles from
tributaries of Tokyo Bay, Japan, were 0.5-53.8 µg/litre. Those in
suspended particles from a wastewater influent were 297-504 µg/litre
and those in the effluent, 0.1-1.22 µg/litre (Takada & Ishiwatari,
1987).
The concentrations of LAS in the estuary of the Krka River,
Croatia, were 420-780 µg/litre near municipal wastewater outlets; 50
m from the wastewater outlets, the concentrations were 7.2 µg/litre
at a depth of 0.5 m and 3.2 µg/litre at a depth of 6 m. The
concentrations in water sampled more than 50 m from the input area
were 1-2 µg/litre. The Krka River estuary was reported to be highly
stratified, with vertical transport of pollutants reduced by the
freshwater-saline boundary. The concentrations of LAS were
negatively correlated with salinity; the maximum concentration,
24 µg/litre, was detected in the surface monolayer. An increase in
the relative abundance of lower homologues of LAS (C10 and C11)
was reported in comparison with the original distribution of
homologues in the wastewater, indicating more rapid depletion of
higher homologues, possibly by biodegradation and fast settling with
particles from sewage (Terzic & Ahel, 1993).
In a comparison of the distribution of homologues of LAS in the
Tama River, Japan, with those established for active substances used
in commercial detergents, the levels of C12 and C13 LAS were
found to decrease over time and those of C10 and C11 to increase
(Hon-Nami & Hanya, 1980a). C11 was the commonest LAS homologue in
river water (Kobuke, 1985; Yoshikawa et al., 1985), and no C13 LAS
were present (Utsunomiya et al., 1980). The average chain length of
LAS in Japanese rivers was C10.9-C11.2 (Nakae et al., 1980;
Yoshimura et al., 1984a; Kobuke, 1985).
Several research groups have confirmed that such changes in
chain length occur during the environmental passage of LAS. In a
study in which the concentration of homologues of LAS was measured
quantitatively by HPLC during activated sludge treatment and lagoon
treatment of wastewater in Spain, the average chain length decreased
from C11.7 in raw material, to C11.3 in the dissolved phase of
raw wastewater, and to C10.3 in the dissolved phase of treated
effluent. A slight increase in average chain length was reported for
the solids compartment in each of these systems, adding to
laboratory findings that the longest homologues adsorb most strongly
to sediment. The reduction in average chain length in the water
compartments was environmentally significant, since shorter
homologues of LAS are less toxic to aquatic organisms. Thus, the
LC50 values for daphnia were higher for shorter homologues (>
20 mg/litre for C11 and 10 mg/litre for C11.7) (Prats et al.,
1993).
The Japanese Soap & Detergent Association (1992) reported a
decrease in LAS concentrations in the Tama River near Tokyo, Japan,
from 2.3 mg/litre in 1967 to 0.2 mg/litre in 1991. The decrease was
attributed to the development of sewage systems along the river:
sewage coverage was 26% in 1974 and 89% in 1990. This information
can be used to estimate concentrations of LAS in developing
countries with inadequate sewage systems but where detergent use is
increasing.
Low levels of LAS were reported in water from the Scheldt River
estuary and in a series of samples from the North Sea (see Table
13). The concentrations in the estuary decreased rapidly from about
0.010-0.012 mg/litre to values below the limit of analytical
detection (0.5 µg/litre) concurrently with an increase in salinity.
The concentrations decreased more rapidly than on the basis of
dilution alone, indicating that removal occurred rapidly. The
authors did not report whether the removal of LAS was related to
adsorption onto settling solids, to biodegradation, or to a
combination of the two. The concentration of LAS in samples taken
offshore was consistently below the limit of detection (Stalmans et
al., 1991).
A5.1.4 Soil and groundwater
The levels of LAS in sludge-amended soil were 0.9-1.3 mg/kg in
German soils used for agriculture. A level of 2.2 mg/kg was found in
the United Kingdom in soil that was used only for the disposal of
sludge (De Henau et al., 1986). MBAS were found at a level of
24.7 mg/kg (14.4-37.5 mg/kg) and LAS at 1.4 mg/kg (0.9-2.2 mg/kg) in
German agricultural soils that had been amended with sludge
(Matthijs & De Henau, 1987). The levels of LAS in soils near the
River Thames, United Kingdom, in 1987 to which sludge had been
applied previously were < 0.2-2.5 mg/kg. Soils that had received an
application of sludge during 1987 had levels of LAS of <
0.2-19.8 mg/kg (Holt et al., 1989).
Levels of 13-47 mg/kg were found on the surface of sludge-
amended soil in the United States in 1979; < 5 mg/kg were found at
a depth of 15-90 cm (Rapaport & Eckhoff, 1990).
A concentration of 22.4 mg/kg LAS was measured in agricultural
soil that had recently been amended with anaerobically digested
sludge. The concentration was 3.1 mg/kg six months after application
of the sludge and 0.7 mg/kg after 12 months (Prats et al., 1993).
HPLC, fluorescence detection, and mass spectrometry were used to
analyse samples of a groundwater plume which originated from an
underground discharge of sewage. It was found that 96% of the LAS
was removed from the aqueous phase during sewage treatment and an
additional 3% during infiltration with groundwater. The
concentrations in ground-water were below the detection limit of
0.01-0.02 mg/litre. The disappearance of LAS during groundwater
infiltration was calculated to follow first-order kinetics. LAS were
detected (by mass spectrometry) at only trace levels in groundwater
sampled 20-500 m down the gradient from the infiltration zone (Field
et al., 1992).
A5.1.5 Drinking-water
The concentration of LAS reported in Dutch tap-water was
0.003 mg/litre; MBAS levels were about three times higher. In
tap-water in the United Kingdom, the concentration of LAS was
0.007 mg/litre; that of MBAS was again three times higher (Waters,
1976). The concentrations of LAS in Italian well-water were below
the analytical limit of detection of 0.0084 mg/litre (Mancini et
al., 1984). LAS were not detected in Japanese drinking-water in the
1970s at a limit of detection of 0.001 mg/litre (Yushi, 1978).
A5.1.6 Biota
The concentrations of LAS in biota are shown in Table 15.
A5.2 Environmental processes that influence concentrations of
linear alkylbenzene sulfonates
A shift towards LAS of lower chain lengths has been reported in
environmental samples in comparison with the distribution of chain
lengths in raw materials. It has also been reported that about 50%
of the total LAS in samples of water is associated with either
suspended particles or dissolved organic matter. Reductions in both
the chain length and the concentration of dissolved LAS will result
in decreased aquatic toxicity (see also section 9).
A5.2.1 Changes in chain length distribution during environmental
removal of linear alkylbenzene sulfonates
The concentrations of LAS and related compounds were measured in
350 samples of water and sediment from the Mississippi River, United
States. Those in surface water were < 0.005 mg/litre. LAS in
sediment had longer chains than those in the overlying water column
(Tabor et al., 1993).
A gradual reduction in the average chain length of homologues
was observed as they passed through a wastewater treatment plant:
untreated wastewater, C12.1; treated effluent, C12; surface
water below a sewage outfall, C11.7 (Castles et al., 1989).
Isomers of C13 LAS have partition coefficients that are typically
one order of magnitude higher than those of the corresponding
isomers of the C12 LAS homologues (Amano et al., 1991).
Table 15. Total body concentrations of linear alkylbenzene sulfonates
in biota in Japan
Organism Year Location Concentration Reference
(mg/kg dry
weight)
Algae 1980-81 River < 1-368 Katsuno et al. (1983)
Pond snail 1979 River 0.4-1.81 Tanaka & Nakanishi
(Sinotaia (1981)
quadratus
histrica)
Gizzard shad 1982 Bay < 1 or < 2 Tokai et al. (1990)
(Konosirus 1983 < 0.1-0.3
punctatus)
A5.2.2 Specification of linear alkylbenzene sulfonates in
surface waters
In most programmes for monitoring LAS in the environment, the
total sample of waste or surface water is analysed, and separate
concentrations of LAS in the fractions of dissolved and suspended
solids are not determined. In a study in which these concentrations
were reported, the mean levels of dissolved LAS were 8.4 mg/litre in
raw wastewater (range, 5.6-11.4 mg/litre) and 5.5 mg/litre in the
suspended solid fraction. In the seven wastewaters studied, an
average of about 65% was present in the filtered (filtration, <
1 µm) 'dissolved' fraction and 35% in the 'solids-associated'
fraction. In treated effluent, 85% of LAS was in the dissolved
fraction and 15% in the solids-associated fraction (Berna et al.,
1993b). In wastewater treatment works, 49-63% of the LAS was in the
dissolved phase and 37-51% in the solids-associated phase (Berna et
al., 1989). In filtered (0.7 µm) wastewater containing LAS at
2.55-2.95 mg/litre, 25-30% LAS was dissolved, and the remaining
70-75% was associated with the solid phase (Cavalli et al., 1991).
The average chain length of homologues of LAS in raw wastewater
was lower in the dissolved phase (C11.2-C11.4) than in the
solids-associated phase (C11.9-C12.0). The authors reported that
39-43% of LAS was present in the dissolved phase and 57-61% in the
solids phase (Prats et al., 1993).
Humic acids extracted from sediments and soils formed strong
association complexes with LAS under environmental conditions, as
observed with fluorescence quenching techniques. The bioavailabilty
of LAS to aquatic organisms is reduced as a result of these
complexes (McAvoy et al., 1993).
A5.3 Estimation of human intake
Human daily intake has been estimated on the assumption that LAS
are taken up from drinking-water and from washing food, vegetables,
dishes, and the skin. The estimates vary from 4.5 to 14.5 mg/day
(Ikeda, 1965; Tokyo Metropolitan Government, 1974; Sterzel, 1992).
The higher figure is based on dubious assumptions about the
concentrations of LAS on vegetables, and the lower value is probably
a more realistic estimate.
The human intake of all anionic surfactants is estimated to be
0.044-0.944 mg/kg per day (Sterzel, 1992), and the maximum daily
intake of ABS, 0.14 mg/kg per day (Ikeda, 1965).
A6. KINETICS
Section summary
LAS are readily absorbed by experimental animals in the
gastrointestinal tract, are distributed throughout the body, and are
extensively metabolized. The parent compound and metabolites are
excreted primarily via the urine and faeces, although there are
marked differences between the isomers in the route of excretion.
The main urinary metabolites identified in rats are
sulfophenylbutanoic acid and sulfophenylpentanoic acid, which are
probably formed through omega-oxidation followed by ß-oxidation of
LAS, although the metabolic pathways in primates may differ.
Although few data are available, it would appear that dermally
applied LAS are not readily absorbed through the skin, although
prolonged contact may compromise the epidermal barrier and permit
more extensive absorption.
A6.1 Absorption, distribution, and excretion
After oral administration of 2 mg/animal of the calcium or
sodium salt of 14C-LAS (chain length, C12) to Wistar rats,
radiolabel was detected in plasma after 0.25 h, reaching maxima at
2 h (0.86 and 1.00 µg/g of the two salts, respectively), and then
decreasing gradually with time; the mean biological half-lives were
calculated to be 10.9 and 10.8 h, respectively. Four hours after
oral administration of the calcium or sodium salt, the concentration
of radiolabel was high in the digestive tract (especially in the
stomach: 22.56 and 31.67 µg/g as the parent compound or metabolites;
and large intestine: 43.24 and 27.26 µg/g) and in the urinary
bladder (34.89 and 16.58 µg/g). The concentrations were also high in
the liver (2.73 and 2.13 µg/g), kidney (1.19 and 1.35 µg/g), testis
(0.08 and 0.11 µg/g), spleen (1.63 and 0.16 µg/g), and lung (0.49
and 0.44 µg/g). At 48 and 168 h, there was little further change.
During the 168-h period after administration, 50% of the radiolabel
on the calcium salt was excreted in urine and 51% in faeces, and 47%
of that on the sodium salt was excreted in urine and 50% in the
faeces (Sunakawa et al., 1979).
Doses of 1 mg per 200 g body weight of two radiolabelled LAS
isomers (chain length, C12) with the benzene sulfonate moieties at
the 2 and 6 positions were administered orally and intravenously to
rats; the same dose was also administered to anaesthetized rats with
bile-duct cannulas by intravenous or intraduodenal injection.
Forty-eight hours after oral or intravenous administration, there
were marked differences in the disposition of the isomers in the
urine and faeces: most of the radiolabel associated with the 2
isomer (75.3%) was in the urine, whereas most of that on the 6
isomer (77.9%) was present in the faeces. After intravenous
administration to bile duct-cannulated rats, 88.6% of the 2 isomer
was recovered in the urine, whereas 83.1% of the 6 isomer was in the
bile. Studies of absorption after intraduodenal administration
showed that both isomers were extensively absorbed within 6 h
(Rennison et al., 1987).
After a dose of 1.2 mg 35S-LAS in aqueous solution was
administered by gavage to bile duct-ligated rats, 89% was absorbed
from the gastro-intestinal tract, as seen by the presence of
radiolabel recovered in urine. Absorption probably occurred mainly
via portal venous blood, since only 1.6% was recovered in the
lymphatic system. When the same dose was administered to bile
duct-cannulated rats, 46% of the radiolabel was recovered in urine,
29% in faeces, and 25% in bile after 90 h. Enterohepatic circulation
was determined in a study in which the bile from one rat was
transmitted to the intestine of another through a cannula; all of
the radioactive LAS excreted in the bile was reabsorbed. In a
separate study, 40-58% of single oral doses of 35S-LAS ranging
from 0.6 to 40.0 mg was excreted in the urine and 39-56% in the
faeces within 72 h of administration (Michael, 1968).
The excretory pattern of 14C-sodium dodecylbenzene sulfonate
was examined in male rats administered a concentration of 1.4 mg/kg
of diet daily for five weeks. The total intake was 1213 µg/rat, of
which 81.8% was excreted during the dosing period, with 52.4% in the
faeces and 29.4% in the urine. After a further week on a normal
diet, however, only 7.8% of the estimated residual amount was found
in excreta. Of a single intraperitoneal injection of 0.385 mg
14C-sodium dodecylbenzene sulfonate/rat (2.26 mg/kg body weight),
84.7% was eliminated within the first 24 h and 94.5% within 10 days
(Lay et al., 1983).
LAS were not detected in the uterus of pregnant ICR mice
administered a single oral dose of 350 mg/kg body weight on day 3 of
gestation (Koizumi et al., 1985).
14C-LAS (chain length, C10-C14, predominantly C11,
C12, and C13) were applied at 250 µg/7.5 cm2 in water to
clipped dorsal skin of rats; the treated area was washed after
15 min, and the animals were restrained from grooming. Most of the
radiolabel was rinsed off, but some of the 14C-LAS
(11 ± 4 µg/cm2) were detected on the treated area; none were
detected in urine or faeces 24 h after the application. In an
accompanying study in vitro, there was no measurable penetration of
14C-LAS (chain length, C12) through isolated human epidermis or
rat skin 24 or 48 h after application (Howes, 1975).
A mixture of 35S-LAS and white petrolatum (29 mg/0.3 ml) was
applied to a 4-cm2 area of the dorsal skin of guinea-pigs, and 24
h after the application about 0.1% of the applied dose was found in
urine and about 0.01% in blood and the main organs. After dermal
application of the same dose to rats and guinea-pigs, the
concentration of 35S in the liver was 9.7 µg/g equivalent of LAS
in rats and about 0.4 µg/g in guinea-pigs (Hasegawa & Sato, 1978).
After a single oral administration of 150 mg/kg 14C-LAS (mean
relative molecular mass, 349) in aqueous solution to rhesus monkeys
(Macaca mulatta), plasma concentrations of radiolabel reached a
maximum equivalent to 41.2 µg/ml at 4 h and then declined over
6-24 h, with a biological half-life of about 6.5 h. The observed
peak plasma concentration of radioactivity (33.6 µg/ml) and the
biological half-life (about 5 h) after seven consecutive daily oral
administrations of 30 mg/kg body weight were similar to those found
after a single administration. The highest concentration of 14C
(238.6 µg/g) was found in the stomach 2 h after the last dose.
Concentrations were also high in the intestinal tract (108 µg/g),
kidney (135.6 µg/g), and liver (64.8 µg/g) and were moderately high
in the lung (19.8 µg/g), pancreas (17.7 µg/g), adrenal glands
(20.6 µg/g), and pituitary gland (17 µg/g). At 24 h, the
concentrations were higher in the intestinal tract (255.4 µg/g) and
liver (10.5 µg/g) than in plasma (2.4 µg/g), whereas those in most
tissues were lower than those in plasma, indicating that there is no
specific accumulation or localization of LAS and their metabolites
in these tissues. After seven subcutaneous doses of 1 mg/kg per day
of 14C-LAS, most of the radiolabel remained in the skin; the
concentration was generally highest at the injection site
(113.96 µg/g). The levels of radiolabel were also high in the
intestinal tract (2.41 µg/g), kidney (1.83 µg/g), lung (2.45 µg/g),
spleen (2.43 µg/g), thyroid (1.24 µg/g), and pituitary (1.00 µg/g)
at 2 h. The concentration in most tissues was generally lower at
4 h, except in the intestinal tract (3.50 µg/g), liver (1.74 µg/g),
and kidney (1.92 µg/g). The high level of radiolabel in the
intestinal tract probably indicates biliary excretion. The average
rates of excretion of radiolabel in urine and faeces during 120 h
after administration of single oral or subcutaneous doses of
14C-LAS to male and female rhesus monkeys are shown in Table 16.
In animals of each sex, radiolabel was excreted primarily in the
urine after either route of administration (Cresswell et al., 1978).
When sodium 35S-dodecylbenzenesulfonate (3.3 mmol/kg body
weight) was administered in the diet to young pigs, at least 35% of
the dose was absorbed through the intestinal tract. After 40 h,
30-40% of the dose had been excreted in urine and > 60% in faeces.
The concentration of radiolabel after 200 h was relatively high in
bristles and bones and low in liver, kidney, and spleen
(quantitative data not presented). After 10 weeks, traceable amounts
of 35S (0.05% of the administered dose) were found in bristles,
bones, skin, lung, and brain (Havermann & Menke, 1959).
Table 16. Excretion of 14C-linear alkyl benzene sulfonates in
rhesus monkeys
Route of administration Sex Concentration (%)
Urine Faeces
Oral (30 mg/kg body weight) Male 68.3 25.9
Female 74.0 20.3
Subcutaneous (1 mg/kg) Male 63.8 12.5
Female 64.3 9.2
From Cresswell et al. (1978); values are average rates of excreted
radioactivity during the 120-h period after a single dose.
A6.2 Biotransformation
The main metabolites isolated from the urine of rats
administered 35S-LAS orally were probably a mixture of sulfophenyl
butanoic (I) and sulfophenyl pentanoic acids (II):
CH3-CH-CH2-COOH CH3-CH-CH2-CH2-COOH
| |
O O
| |
SO3H SO3H
(I) (II)
The material used in the experiment was a mixture of C10-C14 LAS
(mainly C11, C12, and C13). The compounds in this mixture are
probably degraded by omega-oxidation, followed by catabolism through
a ß-oxidation mechanism to form the above metabolites, with
excretion of four or five carbons in the urine (Michael, 1968).
After oral administration of the calcium or sodium salt of
14C-LAS to rats, two metabolites were detected in urine and four
in faeces by thin-layer chromatography. The two urinary and two of
the faecal metabolites were believed to be compounds similar to
metabolites (I) and (II) previously identified by Michael (1968)
(Sunakawa et al., 1979).
Thin-layer chromatography of urine extracts after oral or
sub-cutaneous administration of 14C-LAS to rhesus monkeys showed
only trace amounts of the unchanged compound, and five metabolites
more polar than LAS were detected. These metabolites have not been
identified. Incubation of urine samples with ß-glucuronidase or
sulfatase did not affect the components, which were therefore
probably not present as the corresponding conjugates (Cresswell et
al., 1978).
A7. EFFECTS ON LABORATORY MAMMALS AND IN VITRO TEST SYSTEMS
Section summary
The oral LD50 values for sodium salts of LAS are
404-1470 mg/kg body weight in rats and 1259-2300 mg/kg body weight
in mice. LAS irritate skin and eyes.
Minimal effects, including biochemical alterations and
histopathological changes in the liver, were reported in subchronic
studies in rats administered LAS in the diet or drinking-water at
concentrations equivalent to a dose of about 120 mg/kg body weight
per day. Although ultrastructural changes in liver cells were
observed at lower doses in one study, these changes appeared to be
reversible. Effects have not been seen at similar doses in other
studies, but the organs may have been examined more closely in this
study. Reproductive effects, including decreased pregnancy rate and
litter loss, have been reported in animals administered doses >
300 mg/kg body weight per day. Histopathological and biochemical
changes have been observed following long-term dermal application on
rats of solutions of LAS at concentrations > 5% and after 30 days'
dermal application on guinea-pigs of 60 mg/kg body weight. Repeated
dermal application of solutions containing > 0.3% LAS induced
fetotoxic and reproductive effects, although these doses also
induced maternal toxicity.
The available long-term studies are inadequate to evaluate the
carcinogenic potential of LAS in experimental animals, owing to the
small number of animals used, low or insufficient doses tested, the
absence of a maximal tolerated dose, and limited histopathological
examination. The limited studies available in which animals were
administered LAS orally, however, provide no evidence of
carcinogenicity.
Limited data also indicate that LAS are not genotoxic in vivo
or in vitro.
A7.1 Single exposures
The LD50 values for the sodium and magnesium salts of LAS
given orally, subcutaneously, or intravenously are summarized in
Table 17. Rats appear to be more sensitive than mice to LAS,
regardless of the route of exposure. The LD50 values for LAS given
orally were 1259-3400 mg/kg body weight in mice and 404-1900 mg/kg
body weight in rats. Differences were seen according to the sex,
strain, and age of the animals and the test material.
Table 17. Acute toxicity of linear alkylbenzene sulfonates
Species/ Sex Route LD50a Test materialb Reference
strain (mg/kg
body
weight)
Mouse
NR NR Oral 2170 60% active ingredient Yanagisawa et
al. (1964)
DD M Oral 2300 34.55% solution Tiba (1972)
ddY M Oral 1665 Purified Kobayashi et
ICR-JCL F Oral 1950 Purified al. (1972)
M Oral 1250 Commercial soln, 19.0% Kuwano et al.
F Oral 1540 Commercial soln, 19.0% (1976)
M Oral 1370 Commercial soln, 17.1%
F Oral 1560 Commercial soln, 17.1%
M Oral 2160 99.5% active ingredient Ito et al. (1978)
of C10-C13
F Oral 2250 99.5% active ingredient
of C10-C13
M Oral 2600 Magnesium salt of above
F Oral 3400 Magnesium salt of above
M s.c. 1250 99% active ingredient Ito et al. (1978)
of C10-C13
F s.c. 1400 99% active ingredient
of C10-C13
M s.c. 1529 Magnesium salt of above
F s.c. 1550 Magnesium salt of above
Table 17 (contd)
Species/ Sex Route LD50a Test materialb Reference
strain (mg/kg
body
weight)
ICR-JCL M i.v. 207 99% active ingredient
(contd) of C10-C13
F i.v. 298 99% active ingredient
of C10-C13
M i.v. 98 Magnesium salt of above
F i.v. 151 Magnesium salt of above
NR NR i.v. 120 Yanagisawa et al.
(1964)
Rat
FDRL M,F Oral 650 Nominal chain length, Oser & Morgareidge
C12 (range C9-C15) (1965)
Wistar
6 w M Oral 873 Purified Kobayashi et
6 w F Oral 760 al. (1972)
10 w M Oral 404
10 w F Oral 409
M Oral 1460 99.5% active ingredient Ito et al. (1978)
of C10-C13
F Oral 1470 99.5% active ingredient
of C10-C13
M Oral 1900 Magnesium salt of above
F Oral 1840 Magnesium salt of above
Table 17 (contd)
Species/ Sex Route LD50a Test materialb Reference
strain (mg/kg
body
weight)
CRJ-SD M s.c. 840 99.5% active ingredient
of C10-C13
F s.c. 810 99.5% active ingredient
of C10-C13
M s.c. 710 Magnesium salt of above
F s.c. 730 Magnesium salt of above
M i.v. 119 99.5% active ingredient
of C10-C13
F i.v. 126 99.5% active ingredient
of C10-C13
M i.v. 27.2 Magnesium salt of above
F i.v. 35.0 Magnesium salt of above
NR, not reported; M, male; F, female; s.c., subcutaneous; i.v., intravenous;
w, weeks
a As active ingredient
b Sodium salt, unless specifically indicated
The main clinical signs observed after oral administration of
doses near or greater than the LD50 consisted of reduced voluntary
activity, piloerection, diarrhoea, and weakness. Diarrhoea was more
severe in rats than mice (Kobayashi et al., 1972). Convulsions,
torsion, and paralysis of the hind limbs were also observed in some
of mice (Kobayashi et al., 1972; Kuwano et al., 1976). Death usually
occurred within 24 h of administration. Transient cardiac arrest,
dyspnoea, cyanosis, respiratory collapse, and death occurred during
intravenous injection (Ito et al., 1978).
At autopsy, hyperaemia and haemorrhage of the stomach and
intestine, bloating of the intestine with thinning of its wall, and
congestion of some internal organs were the main macroscopic
findings; histological examination showed congestion and epithelial
degeneration of the gastrointestinal mucosa (Kobayashi et al., 1972;
Kuwano et al., 1976; Ito et al., 1978).
A7.2 Short-term exposure
A7.2.1 Mouse
In a study of the toxicity of a commercial preparation of LAS
(17.1% active ingredient), 44 male and 16 female C57Bl TW mice were
given subcutaneous injections according to the following schedule:
0.02 ml of 1% of the preparation for 10 consecutive days from the
day of birth, 0.04 ml of the same solution for the following 10
days, 0.02 ml of a 10% solution five times over the next 10 days,
and 0.04 ml of the same solution every other day for a further 30 or
60 days. Eight males and six females served as untreated controls.
Epilation and dermatitis usually occurred in animals given
continuous injections of the test material. Adhesions between some
organs, most frequently between the spleen and kidney, were observed
in those receiving injections from the day of birth. Neither the
growth nor the survival of the animals was affected. Although the
weights of the liver, kidney, and spleen were significantly
increased in animals receiving treatment for 60 days,
histopathological examination of the liver, kidney, adrenal glands,
and thyroid by light and electron microscopy showed no evidence of
toxicity (Kikuchi, 1978).
A7.2.2 Rat
A7.2.2.1 Administration in the diet
Groups of five male Wistar rats were fed diets containing LAS
(60% active ingredient; chain length distribution: 10.6% C10, 34.1%
C11, 27.7% C12, 19.0% C13, 8.7% C14) at a concentration
of 0, 0.6, 1.2, or 1.8% (equivalent to 180, 360, or 540 mg/kg body
weight per day) for two and four weeks, and lipids in serum and
liver were analysed. Body weight gain was suppressed in the group
receiving 1.8% at four weeks, and the relative liver weight was
increased at two weeks and thereafter in the groups receiving 1.2
and 1.8%. The levels of triglyceride and total lipids in the serum
had decreased markedly at two weeks in all the experimental groups,
and the levels of phospholipids and cholesterol in the serum had
decreased significantly at two weeks in the groups given 1.2 and
1.8%. These changes were less apparent at four weeks, but
triglyceride, phospholipid, and cholesterol levels in serum were
significantly decreased in the group given 1.8%. Significant
increases in triglyceride levels were seen in the liver after two
weeks in the groups receiving 0.6 and 1.8%, and in cholesterol
levels in the group given 0.6% (Yoneyama & Hiraga, 1977).
Technical-grade sodium LAS (87.9% active ingredient; chain
length distribution: 1.8% C10, 43.2% C11, 32.2% C12, 5.3%
C14, 1.5% C15) were fed to five groups of 10 weanling
Sprague-Dawley rats of each sex at a dietary level of 0, 0.02, 0.1,
or 0.5% (equivalent to 8.8, 44, or 220 mg/kg body weight per day)
for 90 days. No adverse effects were found on survival, growth, food
conversion efficiency, haematological values, urinary analytical
values, or absolute or relative organ weights. There were no gross
or microscopic histological changes attributable to ingestion of the
test material (Kay et al., 1965).
Technical-grade LAS (normal chain length, C12; range,
C9-C15; mean relative molecular mass, 346) were fed to three
groups of weanling FDRL rats, each consisting of 15 males and 15
females, at a dose of 0, 0.05, or 0.25 g/kg body weight per day for
12 weeks. No adverse effects were noted on survival, behaviour,
growth, food conversion efficiency, haematological measurements,
blood chemistry, urine analytical values, organ weights, or gross or
microscopic appearance, except for a slight increase in liver weight
in females given 0.25 g/kg body weight per day (Oser & Morgareidge,
1965).
A diet containing LAS at a concentration of 1.5% (equivalent to
750 mg/kg body weight per day) or a control diet was given to groups
of five male Wistar rats for 2, 4, or 12 weeks. LAS depressed body
weight gain, and the relative liver weight was significantly
increased after two weeks of treatment. The activities of alkaline
phosphatase and glutamate-pyruvate transaminase in serum were
significantly increased at each observation period, and cholesterol
and protein levels were significantly decreased by four weeks. In
the liver, the activities of glucose-6-phosphatase and
glucose-6-phosphate dehydrogenase were decreased, and the activity
of isocitrate dehydrogenase was increased at each observation point.
Enzymatic examination of the renal cortex showed decreased
activities of glucose-6-phosphatase and 5'-nucleotidase at each
observation period, an increase in the activity of lactate
dehydrogenase at 12 weeks, and increased activity of isocitrate
dehydrogenase at 2 and 4 weeks. In the renal medulla, the activity
of Na,K-ATPase was decreased, that of lactate dehydrogenase was
increased at 12 weeks, and that of isocitrate dehydrogenase was
decreased at 2 weeks but increased at 12 weeks (Ikawa et al., 1978).
Groups of five male Wistar rats were given a diet or
drinking-water containing LAS at a concentration of 0.4% (diet:
200 mg/kg body weight per day; drinking-water: 560 mg/kg per day)
for two weeks in order to determine the effects of LAS on the
synthesis of lipids in the liver. Lipids were thus measured in the
liver, and uptake of acetate-1-14C by the lipids was examined.
Decreases in the levels of total lipids and triglyceride were seen
in both groups, but there were no significant changes in
phospholipid or cholesterol levels. Uptake of acetate-1-14C by
lipids in the liver was increased in both groups; uptake of
phospholipids and triglycerides tended to increase, and that of
phospholipids increased significantly in rats given LAS in the diet
(Yoneyama et al., 1978).
A7.2.2.2 Administration by gavage
Groups of 12 male and 12 female Sprague-Dawley rats were given
the magnesium salt of LAS by gavage at a dose of 0, 155, 310, or
620 mg/kg body weight for one month. Body weight gain was depressed
in males and females at 620 mg/kg body weight; one male and two
females at this dose also had diarrhoea and loss of appetite and
subsequently died. Haematological examination revealed significant
decreases in haemoglobin concentration and haematocrit in males at
620 mg/kg body weight. A significant increase in the activity of
alkaline phosphatase and a significant decrease in calcium levels
were seen in males at 310 or 620 mg/kg body weight; and a
significant increase was seen in the activity of glutamate-oxalate
transminase and a significant decrease in protein levels in females
at those doses. Females at all doses had a significant decrease in
calcium levels. At the highest dose, females had a significant
increase in the activity of alkaline phosphatase, a significant
decrease in cholesterol level, and increased weight of the liver,
but the weight of the thymus decreased. The weight of the heart
decreased in females at 310 and 620 mg/kg body weight. Histological
examination of the liver revealed no abnormalities (Ito et al.,
1978).
Groups of 12 male and 12 female Sprague-Dawley rats were given
the sodium salt of LAS (chain length distribution: < 0.1% C9,
10.1% C10, 33.7% C11, 31.0% C12, 25.1% C13) at a dose of 0,
125, 250, or 500 mg/kg body weight by gavage once a day. Diarrhoea
was observed in the group receiving 500 mg/kg, and soft faeces were
observed in the other two groups. Body weight gain was depressed in
males of all groups and in females at 500 mg/kg. Haematological
examination revealed no abnormalities. Serum analysis revealed a
significant increase in the activity of alkaline phosphatase in
males at 500 mg/kg, a significant decrease in calcium levels in
males of all groups, significant increases in the activity of
gluatamate-oxalate transaminase and in blood-urea nitrogen in
females at 500 mg/kg, a significant decrease in calcium level in
females at 250 or 500 mg/kg, and significantly decreased protein and
albumin levels in females of all groups. At 500 mg/kg, the weights
of spleen and heart were significantly decreased in males; in
females, liver weights were increased but the weights of the heart
and thymus were decreased. No histological abnormalities were seen
in the liver (Ito et al., 1978).
A7.2.2.3 Dermal application
Continued, repeated, or extremely high doses of LAS, like other
detergents, compromise the integrity of the skin so that penetration
occurs, causing a variety of anomalies. As the design of the
following two studies was not adequate, the observations are not
considered to be relevant to human risk assessment.
Application of 2 ml of a commercial preparation of LAS (23.4%
active ingredient) to the thoracic skin of six male Wistar rats
resulted in redness and wrinkling of the skin after 24 h. The
redness then increased, the corium was lacerated, and bleeding
occurred. These effects were most severe after five to seven days,
but after a further 10 days the skin began to recover. Six rats died
after 19 days, probably because of the extremely high dose used. The
livers of three rats were examined by electron microscopy after
three and 30 days and the findings compared with those in the
control group. At three days, marked changes were seen in the
components of the liver parenchymal cells, such as separation of the
intracellular space, appearance of dark cells with high electron
density, dysmorphia of mitochondria, extracellular prolapse of
mitochondria, proliferation of rough-surfaced endoplasmic reticulum,
lysosome proliferation, and a decrease in the prevalence of fatty
droplets. At 30 days, many liver parenchymal cells were filled with
abnormally divided and proliferated mitochondria, and an abnormal
increase in smooth-surfaced endoplasmic reticula was noted. There
were no granules of glycogen or fatty droplets. Structures
resembling necrotic cells were also observed (Sakashita et al.,
1974).
A commercial preparation of LAS (23.4% active ingredient) was
applied dermally to male rats (number not given) at a dose of
5 mg/kg body weight active ingredient once a day for 30 days, and
the liver was examined by electron microscopy. Degeneration was seen
in part of the liver, in the form of atrophy and high density.
Intra-mitochondrial deposits and deformation of the Golgi apparatus
were also noted (Sakashita, 1979).
A7.2.2.4 Subcutaneous injection
A commercial preparation of LAS (27% active ingredient) was
given subcutaneously to groups of five male and five female Wistar
rats at a dose of 2 ml/kg body weight per day of a 0, 0.02, 0.2, or
2% solution of the preparation for 25 or 50 days. Rats receiving the
2% solution had reduced body weight gain, increased weights of
liver, kidney, and spleen, a low serum albumin:globulin ratio, low
serum protein, and reduced ornithine aminotransferase activity in
the liver (Hayashi, 1980).
A7.2.3 Guinea-pig
Twelve guinea-pigs were treated daily for 30 days with a
solution of LAS in distilled water equivalent to 60 mg/kg body
weight, which was applied to a 4-cm2 area of clipped dorsal skin.
Twelve controls received acetone at 0.5 ml. The animals were
sacrificed after 30 days, and samples were taken from liver and
kidney and homogenized for determination of enzymes, lipid
peroxidation, glutathione, and protein. The activities of
ß-glucuronidase, gamma-glutamyl transpeptidase, 5-nucleotidase, and
sorbitol dehydrogenase were increased in liver and kidney. Lipid
peroxidation was increased in kidney but not in liver, and the
glutathione content was unchanged in both organs. Extensive fatty
changes were found in hepatic lobules, with dilation of sinusoids;
tubular lesions were found in the kidney, predominantly in the
proximal and distal portions (Mathur et al., 1992).
A7.2.4 Monkey
LAS (chain length, C10-C13) were given to four groups of
three male and three female rhesus monkeys at a daily dose of 0, 30,
150, or 300 mg/kg body weight orally simultaneously with a dose of
0, 0.1, 0.5, or 1.0 mg/kg per day subcutaneously, for 28 days.
Monkeys that received 300 mg/kg orally and 1.0 mg/kg subcutaneously
vomited frequently, usually within 3 h of administration; these
animals and those given 150 mg/kg orally and 0.5 mg/kg
subcutaneously also had an increased frequency of loose or liquid
faeces. Fibrosis at the injection sites was reported in all test
animals, and the incidence and severity were related to dose.
Treatment had no effect on ophthalmoscopic, haematological, or
urinary parameters, on organ weight, or on histopathological
appearance (Heywood et al., 1978).
The studies of short-term exposure to LAS are summarized in
Table 18.
Table 18. Summary of studies of short-term exposure to linear alkylbenzene sulfonates (LAS)
Species, strain, Test material Route Dosage Results Reference
numbers (specification)
per group
Mouse, C57Bl TW LAS (a.i. 17.1%) s.c. 63 or 76 mg/kg Abdominal adhesions, increased Kikuchi (1978)
44 M, 16 bw/day, 60-90 days weights of liver, kidney, and
spleen after 60-day treatment;
no histopathological changes
in liver, kidney, adrenal or thyroid
glands
Rat, Wistar LAS, C10-C14 Diet 0, 0.6, 1.2, 1.8%, Decreased serum triglyceride, Yoneyama & Hiraga
5 M (a.i. 60%) 4 weeks total lipids, phospholipids, and (1977)
cholesterol; increased relative
liver weight at 1.2 and 1.8%;
suppression of body weight gain
at 1.8%
Rat, SD LAS, C10-C15 Diet 0, 0.02, 0.1, 0.5%, No adverse effects Kay et al. (1965)
10 M, 10 F (a.i. 8-9%) 90 days
Rat, FDRL LAS, C9-C15 Diet 0, 0.05, 0.25 g/kg bw Slight increase in liver weight in Oser & Morgareidge
15 M, 15 F (a.i. 39.5%) per day, 12 weeks females at high dose (1965)
Rat, Wistar LAS (NS) Diet 1.5%, 24 weeks Increased activities of serum, Ikawa et al. (1978)
4 M hepatic, and renal enzymes;
depressed body weight gain;
increased relative liver weight
Table 18 (contd)
Species, strain, Test material Route Dosage Results Reference
numbers (specification)
per group
Rat, CRJ-SD LAS, Na, C10-C13 Gavage 125, 250, 500 mg/kg Altered serum enzyme activity Ito et al. (1978)
12 M, 12 F (a.i. 99.5%) bw per day, 1 month and calcium levels at high doses;
decreased serum protein and
albumin levels in all treated
females; decreased spleen and
heart weights in males at highest
dose; increased liver weight and
decreased heart and thymus
weights in females at highest dose;
no histopathological abnormalities
in liver
Rat, CRJ-SD LAS Mg, C10-C13 Gavage 155, 310, 620 mg/kg Altered haemoglobin, haematocrit, Ito et al. (1978)
12 M, 12 F (a.i. 96.9%) bw per day, 1 month serum enzyme activities, calcium
level at high doses; depressed
body weight gain at highest dose;
increased liver weight and
decreased heart and thymus
weights in females at highest dose;
no histopathological abnormalities
in liver
Rat, Wistar LAS detergent Dermal 2 ml/animal Skin irritation; liver parenchymal Sakashita et al.
6 M (a.i. 23.4%) 3.5 × 4.5 cm, 30 days changes with necrotic cells; no (1974)
glycogen granules or fat droplets
Rat, Wistar LAS detergent Dermal 5 mg/kg bw, once/ Degenerative changes in liver Sakashita (1979)
6 M (a.i. 23.4%) day, 30 days
Table 18 (contd)
Species, strain, Test material Route Dosage Results Reference
numbers (specification)
per group
Rat, Wistar LAS detergent s.c. 0, 0.02, 0.2, 2%, Depressed body weight gain; Hayashi (1980)
5 M, 5 F (a.i. 27%) 2 ml/kg bw per day, increased weights of liver, kidney,
50 days and spleen; and altered hepatic
enzyme activities at highest dose
Rat, Wistar LAS Drinking- 0.4%, 2 weeks Decreased hepatic total lipids and Yoneyama et al.
8 M, 8 F (a.i. 60.2%) water triglycerides; increased uptake of (1978)
acetate-1-14C, phospholipids, and
triglycerides
Guinea-pig LAS (NS) Dermal 60 mg/kg bw, Altered hepatic and renal enzyme Mathur et al. (1992)
12 M, 12 F 30 days on 4 cm2 activities; fatty degeneration in
liver; renal tubular lesions
Rhesus monkey LAS C10-C13 Gavage 0.30, 150, 300 mg/kg Vomiting and diarrhoea; no Heywood et al.
3M, 3F (a.i. 20.5%) s.c. 0, 0.1, 0.5, 1.0 mg/kg ophthalmic, haematological or (1978)
bw per day, 28 days urinary changes; no effect on
organ weights; no histopatho-
logical changes
M, male; F, female; a.i., active ingredient; s.c., subcutaneous
A7.3 Long-term exposure; carcinogenicity
A7.3.1 Mouse
A7.3.1.1 Administration in the diet
Groups of eight or nine ICR mice were given diets containing LAS
at a concentration of 0.6 or 1.8% for nine months (corresponding to
intakes of 500 and 1000 mg/kg body weight per day). There was no
reduction in body weight gain at either dose, but the weight of the
liver was increased in both males and females. Significant decreases
were seen in the activities of hepatic lactate dehydrogenase and
renal acid phosphatase in male mice (Yoneyama et al., 1976).
A7.3.1.2 Administration in the drinking-water
Drinking-water containing 100 ppm LAS (corresponding to 20 mg/kg
body weight per day) was supplied to ddy mice (sex and number not
stated) for six months, and they were then allowed to recover for
two months. Mice were killed for electron microscopy of the liver at
one, two, three, and six months and after the two-month recovery
period. Hepatic damage was observed at one and six months,
consisting of the disappearance of the nucleolus, atrophy of the
Golgi apparatus, degranulation of rough-surfaced endoplasmic
reticulum, degeneration of mitochondria, and increased numbers of
primary and secondary lysosomes including autophagic vacuoles with a
myelinated core. In mice examined after the two-month recovery
period, some hepatic damage was seen, which was characterized by
changes in mitochondrial structure and the presence of numerous fat
droplets. Other cellular effects had reversed, indicating that the
liver cells had recovered (Watari et al., 1977). Because an
extremely high dose was used in this study, the observations have
little relevance to human risk.
Groups of eight or nine ICR mice were given water containing LAS
at a concentration of 0.07, 0.2, or 0.6% for nine months,
corresponding to intakes of about 0.1, 0.25, or 0.6 g/kg body weight
per day for males and 0.1, 0.25, or 0.9 g/kg body weight per day for
females. Body weight gain was depressed in males and females at
0.6%, and there were dose-related increases in liver weight in
females in all dose groups. In the group given 0.6% LAS, the
activity of hepatic glutamate-oxalate transaminase was significantly
decreased in males and the activity of renal glucose-6-phosphatase
was decreased in animals of each sex (Yoneyama et al., 1976).
A7.3.2 Rat
A7.3.2.1 Administration in the diet
LAS (98.1% active ingredient; chain length distribution,
C10-C14) were fed to four groups of Charles River weanling rats,
each consisting of 50 males and 50 females, at a dietary level of 0,
0.02, 0.1, or 0.5% (corresponding to 10, 50, or 250 mg/kg body
weight per day) for two years. No adverse effects on growth or feed
conversion efficiency were observed. Five males and females from
each group were killed at 8 and 15 months, and all survivors at 24
months; all animals were necropsied, haematological values were
determined, and tissues were taken for histological examination. No
consistent change was seen that could be considered a toxic
response. Animals that showed significant loss of weight,
development of tumours, or other evidence of abnormalities were also
sacrificed and their tissues preserved for study. The incidences of
tumours and of common incidental diseases were similar in all
dietary groups (Buehler et al., 1971).
Diets containing technical-grade LAS (chain length distribution:
10.6% C10, 34.1% C11, 27.7% C12, 19.0% C13, 8.7% C14; mean
relative molecular mass, 345.8) at a concentration of 0, 0.07, 0.2,
0.6, or 1.8% were given to groups of 10 Wistar rats of each sex for
six months. The group given 1.8% had diarrhoea, markedly depressed
growth, increased caecal weight, and marked degeneration of renal
tubules. The group given 0.6% had slightly depressed growth,
increased caecal weight, increased serum alkaline phosphatase
activity, decreased serum protein, and degeneration of renal
tubules. The group given 0.2% had increased caecal weight and slight
degeneration of renal tubules. The group given 0.07%, corresponding
to about 40 mg/kg body weight per day, showed no effects
attributable to treatment (Yoneyama et al., 1972).
Groups of eight male and eight female Wistar rats were given
diets containing LAS at a concentration of 0, 0.6, or 1.8% for nine
months, corresponding to intakes of 230 or 750 mg/kg body weight per
day for males and 290 or 1900 mg/kg body weight per day for females.
In rats given 1.8% LAS, body weight gain was reduced in both males
and females. Haematological examination revealed a significant
decrease in leukocytes in males at 0.6% and significant decreases in
mean corpuscular volume and mean corpuscular haemoglobin in females
at 1.8%. The activity of glutamate-oxalate transferase and the
levels of cholesterol and albumin in serum were significantly
decreased and the activity of alkaline phosphatase and the levels of
blood-urea nitrogen and cholinesterase were significant increased in
males at 1.8%; females at that dose had a significant decrease in
cholesterol level and a significant increase in alkaline phosphatase
activity. At 0.6%, males had a significant decrease in glucose
level, and females had a significant decrease in the activity of
glutamate-pyruvate transaminase. The caecal weight of male rats and
the liver and caecal weights of female rats at 1.8% were
significantly increased. Enzymatic examination of the liver revealed
dose-related decreases in the activities of glucose-6-phosphate
dehydrogenase and lactate dehydrogenase in male rats. At 1.8%, males
had significantly decreased activities of glucose-6-phosphatase,
glutamate-pyruvate transaminase, and glutamate-oxalate transaminase
and a dose-related decrease in the activity of glucose-6-phosphate
dehydrogenase; females had significantly decreased activities of
glucose-6-phosphatase and glutamate-oxalate transaminase. Enzymatic
examination of the kidneys of females at 1.8% showed significantly
decreased activities of glucose-6-phosphatase, Na,K-ATPase, and
lactate dehydrogenase (Yoneyama et al., 1976).
Groups of 50 male and 50 female Wistar weanling rats were given
diets containing LAS (10.6% C10, 34.1% C11, 27.7% C12, 19.0%
C13, 8.7% C14; mean relative molecular mass, 345.8) at a
concentration of 0, 0.04, 0.16, or 0.6%. In each group, five rats of
each sex were fed for one, three, six, or 12 months, and groups of
15 rats of each sex were fed for 24 months or more. The group fed
0.6% had slightly increased liver and caecal weights, and increased
activity of glutamate-pyruvate transaminase and alkaline phosphatase
in serum. The treatment had no adverse effect on the intake of food,
body weight gain, general condition, mortality, or mean survival. On
the basis of these results, it was concluded that a diet containing
LAS at a concentration of 0.6% (300 mg/kg body weight per day) had
no adverse effects on the rats (Yoneyama et al., 1977).
Groups of 50 male and 50 female Wistar rats were fed LAS
(C10-C14) in the diet at a concentration of 0, 0.04, 0.16, or
0.6% and were then submitted to a detailed histopathological
examination. After one month, proliferation of hepatic cells in the
liver, slight swelling of the renal tubules, and narrowing of the
tubular lumen were found in treated animals. Since these alterations
later disappeared, they were considered to represent adaptation to
the administration of LAS. No histological lesions were seen in the
organs of rats that were fed for 24 months or more that could be
attributed to treatment. Various types of tumour were observed in
both treated and control rats but did not appear to be due to LAS
(Fujii et al., 1977).
A7.3.2.2 Administration in the drinking-water
Groups of eight to nine male and eight to nine female Wistar
rats were given LAS at a concentration of 0, 0.07, 0.2, or 0.6% in
drinking-water for nine months. Body weight gain was suppressed in
males given 0.6%. Haematological examination revealed no significant
change in any of the experimental groups, but a dose-related
decrease in cholesterol level was seen in males. No change in organ
weight was seen that was due to administration of LAS. Significant
decreases in the activities of glutamate-oxalate transaminase and
lactate dehydrogenase were seen in males at 0.2% and a dose-related
increase in the activity of glutamate-oxalate transaminase in
females. A significant decrease in renal Na,K-ATPase was seen in the
group given 0.2%. The dose of 0.07% corresponded to intakes of LAS
of 50 and 120 mg/kg body weight per day in males and females, and
the dose of 0.2% to intakes of 120 and 170 mg/kg body weight per
day, respectively (Yoneyama et al., 1976).
A commercial preparation of LAS (27% active ingredient) was
given to groups of five male Wistar rats in drinking-water at a
concentration of 0, 0.3, 3, 30, or 300 ppm (corresponding to 0.007,
0.07, 0.7, or 7 mg/kg body weight per day) for 60, 124, or 181 days.
Although a reduction in body weight gain, changes in blood
biochemistry, and increased ornithine aminotransferase activity in
the liver were noted in some animals, they were not proportional to
dose or feeding period (Hayashi, 1980).
Groups of 20 male Wistar rats were given water containing LAS
(34.55% commercial solution) at a concentration of 0, 0.01, 0.05, or
0.1% for two years, the highest dose corresponding to an intake of
about 200 mg/kg body weight per day. No changes attributable to the
administration of LAS were seen in terms of growth, mortality, the
weights of major organs, or histopathological appearance (Tiba,
1972).
A group consisting of 62 male and 62 female Wistar rats was
given drinking-water containing LAS (mean relative molecular mass,
348; 38.74% active ingredient) at a concentration of 0.1%
(corresponding to 140 mg/kg body weight per day), and a control
group of 37 male and 37 females was given normal drinking-water.
Five to 12 rats in the experimental group and three to 12 rats in
the control group were killed at 3, 6, 12, and 18 months, and all
surviving animals were killed at 24-26 months. Administration of LAS
had no effect on the intake of water, mortality, body weight gain,
or general condition. Histopathological examination revealed
atrophy; fatty changes were found in hepatic cells in treated
animals at six months, when there were also significant increases in
the activities of glutamate-oxalate and glutamate-pyruvate
transaminases and in the level of bilirubin. LAS had no effect on
haematological parameters (Endo et al., 1980).
A group of 60 male and 60 female rats (strain not specified)
received drinking-water containing 0.01% of a preparation containing
51% LAS for 100 weeks; a similar group was untreated. No detrimental
effects on body weight and no pathological effects, including
tumours, were reported (Bornmann et al., 1963).
A7.3.2.3 Administration by gavage
Groups of 20 male and 20 female Sprague-Dawley rats were given a
solution of a magnesium salt of LAS at doses of 10, 75, 150, or
300 mg/kg body weight per day by gavage for six months. Body weight
gain was suppressed, and slight decreases were observed in serum
protein, albumin, and calcium ion level, but the changes were within
the physiological range (Ito et al., 1978).
A7.3.2.4 Dermal application
A dose of 0.1 ml/kg body weight of a 0.5, 1.0, or 5.0% solution
of magnesium LAS (in 3% polyethylene glycol) was applied to the
backs of 20 male and 20 female Sprague-Dawley rats six times a week
for six months. Slight redness at the application site was observed
transiently in males and occasionally in females at 5%. Body weight
was slightly suppressed in males at that dose, and one male in the
control group and one at 5.0% died of unknown causes. Treatment had
no definite effect in terms of food conversion efficiency, urinary,
haematological, serum biochemistry, or histopathological findings,
or organ weights (Ito et al., 1978). No systemic toxicity was
reported in this study. Sakashita et al. (1974) and Sakashita (1979)
(see section 7.2.2.3) may have obtained positive results because
they used a shorter period of exposure, during which skin integrity
may have been compromised, resulting in absorption of the
preparation of LAS through the skin to produce systemic effects.
LAS (19.7% active ingredient) were applied to the dorsal skin of
SLC-Wistar rats three times per week at a dose of 0.005, 0.025, or
0.125 ml/rat (equivalent to 1, 5, or 25 mg/rat) for 24 months. A
dose of 0.025 ml of an LAS-based detergent containing 19.9% LAS
(equivalent to 5 mg LAS per rat) and distilled water was given to
controls. Each application was washed from the skin with warm water
after 24 h. Treatment had no effect on organ weights or
histopathological appearance, and there was no evidence of toxicity
or carcinogenicity (Taniguchi et al., 1978).
Long-term studies of exposure to and the carcinogenicity of LAS
are summarized in Table 19.
A7.4 Skin and eye irritation; sensitization
The potential of LAS to irritate the skin depends on the
concentration applied. On the basis of the criteria of the European
Commission and the OECD test guideline, LAS were classified as
irritating to the skin at concentrations above 20% (European
Committee of Organic Surfactants and Their Intermediates, 1990).
Table 19. Summary of studies of long-term exposure to linear alkylbenzene sulfonates (LAS)
Species, strain, Test material Route Dosage Results Reference
numbers (specification)
per group
Mouse, SLC-ICR LAS (a.i. 60%) Diet 0, 0.6, 1.8%, Increased liver weight; Yoneyama et al.
8-9 M, 8-9 F 9 months decreased hepatic and renal (1976)
enzyme activities in males
Mouse, ddy (NR) LAS (NS) Drinking- 20 mg/kg bw per Degenerative changes in liver, Watari et al. (1977)
water day, 6 months with partial recovery after
end of treatment
Mouse, ICR LAS (a.i. 60%) Drinking- 0, 0.07, 0.2, 0.6, Depressed body weight gain at Yoneyama et al.
8-9 M, 8-9 F water 1.8%, 9 months high dose; dose-related increase (1976)
in liver weight in all treated
females; changes in hepatic
enzyme activities at high dose
Rat, Wistar LAS, C10-C14 Diet 0, 0.07, 0.2, 0.6, Dose-related depression of Yoneyama et al.
10 M, 10 F 1.8%, 6 months growth, caecal enlargement, (1972)
and renal tubular degeneration
at > 0.07%
Rat, Wistar LAS (a.i. 60%) Diet 0, 0.6, 1.8%, Depressed body weight gain Yoneyama et al.
8 M, 8 F 9 months at high dose; changes in (1976)
haematological parameters, in serum
and hepatic enzyme activities,
and in cholesterol levels at both
doses; changes in renal enzyme
activities in females at high dose
Table 19 (contd)
Species, strain, Test material Route Dosage Results Reference
numbers (specification)
per group
Rat, Wistar LAS (a.i. 60%) Drinking- 0, 0.07, 0.2, 0.6%, Depressed body weight gain in Yoneyama et al.
8-9 M, 8-9 F water 9 months males at high dose; no changes (1976)
in haematological parameters or
organ weight; changes in serum
and renal enzyme activities at 0.2%
Rat, Wistar LAS, C10-C14 Diet 0, 0.04, 0.16, 0.6%, Slight increase in liver and Yoneyama et al.
50 M, 50 F (a.i. 60%) 24 months caecal weights and changes in (1977)
serum enzym activities at high
dose; no effect on body weight
gain
Rat, Charles River LAS, C10-C14 Diet 0, 0.02, 0.1, 0.5%, No treatment-related effects Buehler et al.
50 M, 50 F (a.i98.1%) 2 years (1971)
Rat, Wistar LAS, C10-C14 Diet 0, 0.04, 0.16, 0.6%, Transient changes in liver and Fujii et al. (1977)
50 M, 50 F (a.i. 60%) 2 years kidney; no treatment-related
histopathological abnormalities
at end of study
Rat, SD LAS Mg, C10-C13 Gavage 75, 150, 300 mg/kg Depressed body weight gain; no Ito et al. (1978)
20 M, 20 F (a.i. 96.9%) bw per day, 6 significant adverse effects
months
Rat, Wistar, 5 M LAS detergent Drinking- 0, 0.3, 3, 30, 300 Depressed body weight gain and Hayashi (1980)
(a.i. 27%) water ppm, 181 days changes in blood biochemistry
and liver enzyme activity considered
not to be related to treatment
Table 19 (contd)
Species, strain, Test material Route Dosage Results Reference
numbers (specification)
per group
Rat, Wistar, 20 M LAS (a.i. 34.55%) Drinking- 0, 0.01, 0.05, 0.1%, No adverse effects Tiba (1972)
water 2 years
Rat, Wistar LAS (a.i. 38.74%) Drinking- 0, 0.1%, 26 months Fatty changes and atrophy in Endo et al. (1980)
62 M, 62 F water liver; changes in hepatic enzyme
activities; no effect on body
weight gain
Rat LAS (Marlon Drinking- 0, 0.01%, No adverse effects Bornmann et al.
60 M, 60 F BW 2043) water 100 weeks (1963)
Rat, SD LAS Mg, C10-C13 Dermal 0.5, 1.0, 5% in Slight reduction in body weight Ito et al. (1978)
20 M, 20 F (a.i. 96.9%) polyethylene glycol, gain of males at high dose; no
6 months other adverse effects
Rat, SLC-Wistar LAS (a.i. 19.7%) Dermal 0, 6.7, 33.3, 167.0 No adverse effects Taniguchi et al.
25 M, 25 F mg/kg bw, 3 × per (1978)
week, 2 years
Rat, SLC-Wistar LAS detergent Dermal 0, 33.3 mg/kg bw No adverse effects Taniguchi et al.
25 M, 25 F (a.i. 19.9%) 3 × per week, 2 years (1978)
M, male; F, female; NS, not specified; a.i., active ingredient; SD, Sprague-Dawley
A7.4.1 Studies of skin
Solutions of LAS (chain length distribution, C10-C13;
purity, 99.9%) were applied to the backs of groups of three male
Wistar rats at a rate of 0.5 g of a 20 or 30% solution once a day
for 15 days. On the sixteenth day of the experiment, the skin at the
application site and the tissues of the tongue and oral mucosa (to
examine the effects of licking) of the rats that received 30% were
examined histologically. Body weight gain was reduced in the group
exposed to 20%, and body weight was decreased in animals exposed to
30%. An infiltrating, yellow-red brown crust was observed after two
to three days at 20% and after one to two days at 30%; at four to
six days, the crust was abraded, and erosion was observed.
Histological examination of the application site revealed severe
necrosis of the region, from the epidermis cuticle to the upper
layer of the dermis, severe infiltration of leukocytes in the
necrotic site, diffuse inflammatory cell infiltration of all of the
layers of the corium, and swelling of collagenous fibres in the
dermis. Histological examination of the tongue showed no changes,
but examination of the oral mucosa revealed atrophy and slight
degeneration of the epithelium (Sadai & Mizuno, 1972).
Some batches of a paste of LAS (volume not stated) induced weak
to moderate sensitization in guinea-pig skin at induction
concentrations of 2-100% and challenge concentrations of 1-2%. A
prototype liquid laundry detergent (10% LAS) induced sensitization
at a challenge concentration of 1% (0.1% as LAS) (Nusair et al.,
1988).
The biochemical and pathomorphological effects of LAS on the
skin of four female albino CDRI guinea-pigs were investigated by
shaving the abdominal skin and immersing the animals up to the neck
in a 1% aqueous solution of neutralized LAS for 90 min daily for
seven consecutive days. A control group was immersed in water
according to the same schedule. After each immersion, the animals
were washed and their skin dried. The animals were killed after
seven days, and skin samples were taken. The skin of guinea-pigs
exposed to the solution of LAS had increased activity of histidine
decarboxylase, decreased sulfhydryl groups and histamine, and
decreased activity of lactic dehydrogenase. It appeared to be
shrunken, with thinner layers of dermis and epidermis than controls.
There were also areas of scarring in the epidermis and ridging of
epidermis and dermis (Misra et al., 1989a).
A7.4.2 Studies of the eye
A volume of 0.1 ml of a solution of LAS (relative molecular
mass, 346.5) at five concentrations ranging from 0.01 to 1.0% was
instilled into the eyes of rabbits (13 per group). The rabbits were
observed for 24 h after application. The group receiving 0.01% had
no abnormalities, but that given 0.05% had slight congestion.
Concentrations of 0.5% and more induced marked reactions, such as
severe congestion and oedema, increased secretion, opacity of the
cornea, and disappearance of the corneal reflex (Oba et al., 1968a).
Solutions of LAS (chain length distribution, C10-C14; 80.9%
C11-C13) at six concentrations ranging from 0.01 to 5.0% were
instilled into the eyes of rabbits (three per group). The rabbits
were observed for 168 h after application. The group given 0.01% had
no reaction, but within 2 h those given 0.05% had slight congestion
and those at 0.1% had considerable congestion or oedema, which had
disappeared by 24 h. Animals given 0.5% or more had marked
reactions, such as severe congestion and oedema, increased
secretion, opacity of the cornea, and disappearance of the corneal
reflex, for 24 h but then tended to recover; the signs had
disappeared completely within 120 h (Iimori et al., 1972).
A7.5 Reproductive toxicity, embryotoxicity, and teratogenicity
The reproductive toxicity of LAS and formulations of LAS has
been evaluated in studies by oral (gavage, diet, drinking-water),
dermal (skin painting), and parenteral (subcutaneous)
administration. Similar effects were seen, regardless of the route
of application. The studies had a number of deficiencies, however,
which are summarized below.
In some studies, widely separated dose levels were used (Palmer
et al., 1975a; Takahashi et al., 1975; Tiba et al., 1976; Hamano et
al., 1976), so that it is difficult to assess dose-response
relationships and to interpret the results. Some of the studies
included only one dose (Bornmann et al., 1963; Sato et al., 1972;
Endo et al., 1980) and some two (Iimori et al., 1973; Nolen et al.,
1975; Takahashi et al., 1975; Hamano et al., 1976; Tiba et al.,
1976). The studies done on formulations are difficult to interpret,
as the effects seen may have been due to another component. In some
cases, the details of the formulation are not given, so that the
dose of LAS is also unknown. Certain studies of dermal exposure
(Sato et al., 1972; Masuda et al., 1973, 1974; Palmer et al., 1975a;
Nishimura, 1976; Daly et al., 1980) involved levels that compromised
the integrity of the skin and caused overt toxicity.
The teratogenic effects of some commercial formulations of LAS
reported by Mikami and co-workers (1969), mainly in mice, were not
reproduced in other studies. A number of studies indicated that LAS
have some reproductive toxicity, but the effects were seen only at
doses that caused maternal toxicity. No teratogenic effects were
observed. These studies are summarized in Tables 20-22.
Table 20. Studies of the reproductive toxicity and teratogenicity of linear alkylbenzene sulfonates (LAS) and formulations of LAS,
administered orally
Route Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
LAS
Diet Charles River 14, 70, 350 84 Combined study of reproduction Buehler et al. (1971)
rats (20) (0.02, 0.1, 0.5%) and teratogenicity (three generations);
no effects attributable to LAS
Diet SD rats (16) 78, 780 (0.1, 1.0%) 0-20 No abnormalities at either dose; few Tiba et al. (1976)
offspring at high dose
Gavage ICR mice (NS) 300, 600 6,8,10 High incidence of cleft palate and Mikami et al. (1969)
exencephaly in fetuses at high dose
Gavage ICR mice (14) 40, 400 (0.4, 4.0%) 0-6 No effects at low dose; reduced weight Takahashi et al.
7-13 gain and pregnancy rate at high dose (1975)
Gavage ICR mice (25-33) 10, 100, 300 6-15 Reduced weight gain at all levels, Shiobara & Imahori
particularly at highest dose; two (1976)
dams died at highest dose; all
fetuses of one dam died in utero;
decreased body weight and delayed
ossification in living fetuses but no
increase in incidence of malformations
Table 20 (contd)
Route Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
LAS (contd).
Gavage ICR mice 14, 20, 350 1-3 No effect on implantation rate at Koizumi et al. (1985)
any dose
Gavage CD rats (20) 0.2, 2.0, 300, 600 6-15, rats No effects on any species at two lower Palmer et al. (1975a)
CD-1 mice (20) and mice doses
NZW rabbits (13) 6-18, Rats: reduced weight gain and one
rabbits death at highest dose
Mice: reduced weight gain, seven
deaths, and four litter losses at 300 mg/kg
bw per day; 18 deaths, one litter loss
and one non-pregnancy at 600 mg/kg
bw per day
Rabbits: reduced weight gain, 11 deaths,
two litter losses at 300 mg/kg bw per
day; all animals died at highest dose
Gavage CD rats (30) 125, 500, 2000 6-15 Two-generation study of reproductive Robinson &
and developmental toxicity; delayed Schroeder (1992)
ossification significant at highest dose,
slight at middle dose; no reproductive
or developmental toxicity
Table 20 (contd)
Route Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
LAS (contd).
Drinking- Charles River 7 (0.01%) Three-generation study of fertility; no Bornmann et al.
water rats (10) teratogenic effects (1963)
Drinking- Wistar rats (20) 70 (0.1%) Four-generation study of reproductive Endo et al. (1980)
water toxicity; no effects attributable to LAS
Drinking- Wistar rats (20) 383 mg/rat (0.1%) 6-15 No effects in rats; rabbits had reduced Endo et al. (1980)
water NZW rabbit (11) 3030 mg/rabbit 6-18 weight gain and delayed ossification
(0.1%) but no malformations
17% LAS, 7% alcohol ethoxylate sulfate
Gavage CD rats (20) 0.8, 8, 1,200, 2400 6-15 No increase in major malformations Palmer et al. (1975a)
CD-1 mice (20) 1.064, 10.64, 6-15 or significant changes in anomalies
1600, 320
NZW rabbits (13) 0.8, 8, 1200, 2400 6-18
45% LAS
Diet CD rats (25) 80, 400, 800 6-15 No treatment-related effects on Nolen et al. (1975)
(0.1, 0.5, 1.0%) reproduction or embryonic
development
Table 20 (contd)
Route Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
1% LAS
Gavage ICR mice (18-23) 800, 1200, 1500, 6-15 No increase in fetal malformations; Yamamoto et al.
3000 decreased body weight and delayed (1976)
ossification at 1200 mg/kg bw
19% LAS
Gavage IRC mice (9-13) 125, 4000 6 No effect on fetal viability or Hamano et al.
development (1976)
NS, not specified
Table 21. Studies of the reproductive toxicity and teratogenicity of linear alkylbenzene sulfonates (LAS) and formulations of
LAS, administered dermally
Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
LAS
CD rats (20) 0.6, 6.0, 60 1-15 Slight reduction in body weight gain at Palmer et al.
(0.03, 0.3, 3.0%) highest dose; no effect on litter parameters (1975a)
at any dose; no evidence of malformations
CD-1 mice (20) 5, 50 , 500 2-13 Reduced body weight gain, fewer pregnancies,
(0.03, 0.3, 3.0%) and total litter loss at highest dose; no
malformations
NZW rabbits (13) 0.9, 9, 90 1-16 Marked reduction in body weight gain, fewer
(0.03, 0.3, 3.0%) pregnancies, and two litter losses at highest
dose; reduced body weight gain at 9 mg/kg bw
per day; no malformations
Wistar rats (20) 20, 100, 400 0-20 Reduced body weight gain, decreased Nishimura (1976)
(1, 5, 20%) pregnancy rates and delayed ossification
at highest dose; no effects at lower doses
Wistar rats (20) 20, 100, 400 0-20 Irritation at site and reduced body weight Daly et al. (1980)
(1, 5, 20%); gain at two higher doses; no change in fetal
rinse-off parameters at any level
Table 21 (contd)
Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
Wistar rats (contd)
0.1, 2, 10 0-20 No change in fetal parameters at any level
(0.05, 0.1, 0.5%);
leave on
ddy/s mice (16) 110 (2.22%) 0-13 No abnormalities in dams or fetuses Sato et al. (1972)
ddy mice (4-10) 0.084, 0.84, 8.4 2-14 No fetal or reproductive effects Masuda et al.
(0.017, 0.17, 1.7%) (1973, 1974)
ICR mice 4.2, 8.4, 12.0, 16.5 1-13 Delayed ossification at two highest
(25-30) (0.85, 1.7, 2.55, 3.4%) doses
ICR mice 15, 150, 1500 6-15 Clear decrease in pregnancy rate and Imahori et al.
(27-28) (0.03, 0.3, 3.0%) decrease in fetal weight at highest dose; (1976)
no increase in malformations in fetus
17% LAS, 7% ethanol, 15% urea
ICR mice 2.5, 25, 75 1-13 Decrease in pregnancy rate at Inoue & Masuda
(11-20) (0.5, 5, 15%) highest dose; no other effects (1976)
Table 21 (contd)
Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
16.3% LAS
ICR mice 25, 50, 100 0-13 Reduced pregnancy rate and Nakahara et al.
(17-50) (5, 10, 20%) some total litter losses at (1976)
highest dose
Unknown formulation
ddy/s mice 65 (15%) 0-13 Decreased body weight gain, Sato et al. (1972)
(21) decreased pregnancy rate,
decreased fetal weight, and delayed
ossification
Unknown formulation
IRC mice 75, 100 0-12 Decreased pregnancy rates Iimori et al. (1973)
(27-39) (15, 20%) at both levels
Unknown formulation
IRC mice 30, 65, 85, 0-13 Decreased pregnancy rates at all Takahashi et al.
(15-19) 100, 125 doses; decreased fetal body (1975)
(13.0, 17.0, weight; delayed ossification at all
20.0, 25.0%) doses except 65 mg/kg bw per day
Table 22. Studies of the reproductive toxicity and teratogenicity of linear alkylbenzene sulfonates (LAS) and LAS formulations,
administered subcutaneously
Species (no. of Dose (mg/kg Length of Comments and results Reference
animals/group) bw per day) treatment
(days)
LAS
ICR mice 0.4, 2.0, 10% 7-13 No significant effects on dams; high Masuda & Inoue (1974)
(21-24) incidence of skeletal variations and
delayed ossification, not dose-related;
no abnormalities
ICR mice 20, 200 0-3 Irritation at injection site and reduced Takahashi et al. (1975)
(12-19) (0.35, 1.00%) 8-11 pregnancy rate at highest dose; no
malformations or anomalies
17% LAS, 7% ethanol, 15% urea
CR mice 30, 150 7-13 No increase in major malformations Inoue & Masuda (1976)
(16-17) 0-13 or minor anomalies; increase in
implantations at high dose given on
days 0-13
A7.6 Mutagenicity and related end-points
A7.6.1 Studies in vitro
Assays for mutagenicity were performed in vitro with two
commercial products containing 17.1 and 19% LAS, either undiluted or
diluted 10 and 100 times (Oda et al., 1977), 99.5% pure LAS (Fujita
et al., 1977), 95.5% pure sodium salt, or 96.2% pure calcium salt
(Inoue & Sunakawa, 1979), using Bacillus subtilis H17 (rec+) and
M45 (rec-), Salmonella typhimurium TA98 and TA100 (including
a metabolic activation system), and Escherichia coli WP2 uvrA.
All of the assays gave negative results. LAS 99.5% pure (Fujita et
al., 1977) were also tested in S. typhimurium TA1535 and TA1537,
again with negative results. Thesodium and calcium salts in the
presence of various liver homogenates (Sunakawa et al., 1981) and a
22.2% solution of LAS (C10-C14, 10-200 µg/plate) (Inoue et al.,
1980) were tested in S. typhimurium TA98 and TA100. No
mutagenicity was seen.
A7.6.2 Studies in vivo
Groups of male ICR:JCL mice were given LAS at a dose of 200,
400, and 800 mg/kg body weight per day by gavage for five days and
were killed 6 h after the final administration for examination of
chromosomal aberrations in bone-marrow cells. One commercial
preparation containing 19.0% LAS was also given, at a dose of 800,
1600, or 3200 mg/kg body weight, and another containing 17.1% LAS at
a dose of 1000, 2000, or 4000 mg/kg body weight once only by gavage.
The highest doses were 50% of the respective LD50 values. Bone
marrow was examined 6, 24 and 48 h after administration. There was
no significant difference between any of the groups given LAS and
the negative control group in the incidence of chromosomal
aberrations. Mitomycin C, used as a positive control at 5 mg/kg body
weight, induced severe chromosomal aberrations (Inoue et al., 1977).
Groups of five male Wistar rats, Sprague-Dawley rats, and ICR
mice were given a diet containing 0.9% LAS for nine months. The
equivalent doses were 450 mg/kg body weight per day in rats and
1170 mg/kg body weight per day in mice. There were no significant
differences in the incidence of chromosomal aberrations between the
experimental and control groups (Masubuchi et al., 1976).
After LAS (C10-C15) were fed to groups of six male and six
female Colworth/Wistar rats in the diet at concentrations of 0.56 or
1.13%, equivalent to 280 or 565 mg/kg body weight per day, for 90
days, no alteratuons were seen in chromosomes in bone marrow (Hope,
1977).
In three male ddY mice given LAS at 100 mg/kg body weight by
intraperitoneal injection, there was no differences between the
treated animals and a control group in the incidence of
polychromatic erythrocytes with micronuclei in bone-marrow cells
(Kishi et al., 1984).
An assay to detect dominant lethal mutations was performed in
seven male ICR:JCL mice given a diet containing 0.6% LAS at
300 mg/kg body weight per day for nine months. Each of the male mice
was then mated with two female mice that had not been given LAS, and
11 of the 14 females became pregnant. The pregnant mice were
laparotomized on day 13 of gestation to determine the numbers of
luteal bodies, implantations, surviving fetuses, and dead fetuses.
There were no significant differences in fertility, mortality of ova
and embryos, the number of surviving fetuses, or the index of
dominant lethal induction (Roehrborn) between the experimental and
control groups (Masubuchi et al., 1976).
LAS were administered as a single oral dose of 2 mg to pregnant
ICR mice on day 3 of gestation; on day 17 of gestation, each animal
received a subcutaneous dose of 1, 2, or 10 mg/mouse and was killed
24 h later. There was no difference among treated groups in the
incidence of polychromatic erythrocytes with micronuclei in maternal
bone marrow or fetal liver or blood. No mutagenic effect was found
in any of the groups (Koizumi et al., 1985).
A7.7 Special studies
A7.7.1 Studies in vitro
The haemolytic action of LAS was investigated by mixing red
blood cells from rabbits with solutions of LAS at concentrations of
1-1000 mg/litre at 38°C for 30 min. Haemolysis occurred at
concentrations > 5 mg/litre (Yanagisawa et al., 1964). Red blood
cells from rabbits were mixed with solutions of various
concentrations of LAS (relative molecular mass, 346.5) at room
temperature for 3 h. The 50% haemolytic concentration of LAS was
9 mg/litre (Oba et al., 1968a).
Purified LAS at various concentrations were added to 10 µl of
normal plasma obtained from male rats, and prothrombin time was
determined. Prothrombin time was prolonged; the 50% inhibitory
concentration was about 0.6 mmol/litre. When LAS at various
concentrations were added to a mixture of 1% fibrinogen and
thrombin, the time of formation of a mass of fibrin was prolonged by
inhibition of thrombin activity. The 50% inhibitory concentration
was about 0.05 mmol/litre (Takahashi et al., 1974).
LAS influenced the thermal denaturation and decreased the
fluorescence profile of bovine serum albumin in vitro, indicating
protein-LAS interaction (Javed et al., 1988).
Eggs from female B6C3F1 mice were fertilized in vitro and
incubated in culture medium containing LAS at concentrations between
0.015 and 0.03%; eggs grown in culture medium without LAS served as
controls. Eggs exposed for 1 h, washed, and then cultured for five
days developed normally to the blastocyst stage when the
concentration of LAS was less than 0.025%; at concentrations higher
than 0.03%, the eggs did not develop beyond the one-cell stage. With
continuous exposure to LAS for five days, a concentration of 0.01%
slightly impaired development to the blastocyst stage, and 0.025%
prevented development to the one-cell stage (Samejima, 1991).
LAS with a chain length distribution of C10-C14 did not
induce transformation of cryopreserved primary cultures of Syrian
golden hamster embryo cells in vitro (Inoue et al., 1979, 1980).
A7.7.2 Biochemical effects
The levels of amylase, alkaline phosphatase, glutamate-oxalate
transaminase, and glutamate-pyruvate transaminase and of the
electrolytes Ca, P, and Mg in serum were determined up to 24 h after
a single oral administration of 2, 5, 50, or 100 mg/kg body weight
of LAS (60% active ingredient) or dermal application of 5 ml of a 1,
5, 10, or 20% solution of LAS to rabbits (number not stated). The
levels of total Ca, Ca2+, Mg, and P were generally lower after
either type of administration than before. Although there was no
definite trend, the activities of the enzymes tended to decrease
regardless of the route of the administration or the dose
(Yanagisawa et al., 1964).
Groups of three male mice were given an intraperitoneal
injection of 0.3 g/kg body weight of LAS (C14) in order to study
the effects on the formation of methaemoglobin, determined 0.5, 1,
and 2 h afterinjection of LAS. The level of methaemoglobin in the
experimental groups was not significantly greater than that in the
control group at any time (Tamura & Ogura, 1969).
The effects of LAS (sodium dodecylbenzenesulfonate) on fasting
blood glucose level and glucose tolerance curves were investigated
in 40 male and 50 female albino rats pretreated with 0.25 g/kg body
weight per day of LAS for three months. At the end of this period,
the rats were divided into four groups and given distilled water,
6.1 g/kg body weight of glucose, 0.94 g/kg body weight of LAS, or
6.1 g/kg body weight of glucose plus 0.94 g/kg body weight of LAS by
gavage. Blood glucose was then estimated at 30-min intervals.
Administration of LAS in conjunction with glucose resulted in higher
initial levels of blood glucose in male rats and persistently higher
levels in females than did administration of glucose alone. Females
in control and pretreated groups generally had higher blood glucose
levels in response to administration of glucose or LAS plus glucose
than did male rats (Antal, 1972).
A8. EFFECTS ON HUMANS
Section summary
Human skin can tolerate contact with solutions of up to 1% LAS
for 24 h with only mild irritation. Like other surfactants, LAS can
delipidate the skin surface, elute natural moisturizing factor,
denature the proteins of the outer epidermal layer, and increase
permeability and swelling of the outer layer. LAS do not induce skin
sensitization in humans, and there is no conclusive evidence that
they induce eczema. No serious injuries or fatalities have been
reported following accidental ingestion of LAS-containing surfactant
preparations.
A8.1 Exposure of the general population
Surface-active agents are used in shampoos, dish-washing
products, household cleaners, laundry detergents, and other
applications such as industrial cleaners. LAS are major components
of such products. In general, the concentration of nonionic and
ionic surfactants is 10-20%.
A8.2 Clinical studies
A8.2.1 Skin irritation and sensitization
LAS are mildly to moderately irritating to human skin, depending
on the concentration. There is no evidence that they sensitize the
skin in humans.
The relative intensity of skin roughness induced on the surface
of the forearms of volunteers (a circulation method) due to contact
with LAS of different alkyl chain lengths (C8, C10, C11-C16)
was characterized mainly by gross visible changes. C12 LAS
produced more skin roughening than LAS with longer or shorter alkyl
chains. The degree of skin roughening in vivo correlated with the
extent of protein denaturation measured in vitro (Imokawa et al.,
1975a).
Primary skin irritation induced by an LAS formulation (average
chain length, C12; relative molecular mass, 346.5), by
alpha-olefin sulfonates (AOS) (27% C15, 25% C16, 28% C17, 8%
C18; relative molecular mass, 338.5), and by alkyl sulfates (AS)
(C12; relative molecular mass, 346.5) was compared in a 24-h
closed-patch test on the forearms of seven male volunteers. A 1%
aqueous solution (pH 6.8) of each substance was used, and the
relative intensity of skin irritation was scored by grading
erythema, fissuring, and scales. The average score for LAS was
similar to that for AOS but significantly lower than that for AS
( p < 0.05) (Oba et al., 1968a).
In another comparison, the intensity of skin irritation induced
by 1% aqueous solutions of LAS (C10-C13), AOS (C14, C16,
C18), and the sodium salt of AS (C12-C15) was studied in a
24-h closed-patch test on the forearm and in a test in which the
substance was dripped onto the interdigital surface for 40 min once
daily for two consecutive days at a rate of 1.2-1.5 ml/min. Skin
reactions were scored by grading erythema in the patch test and by
grading scaling in the drip test. In the patch test, the score for
LAS was similar to that for AOS but significantly lower than that
for AS. In the drip test, the score for LAS was similar to that for
AS but higher than that for AOS (Sadai et al., 1979).
Repeated patch tests with LAS at aqueous concentrations of 0.05
and 0.2% produced mild to moderate primary irritation. In a study on
the sensitization potential of LAS for human skin, a 0.1% aqueous
preparation caused no sensitization in 86 subjects (Procter & Gamble
Co., unpublished data).
No skin sensitization was seen in 2294 volunteers exposed to LAS
or in 17 887 exposed to formulations of LAS (Nusair et al., 1988).
A8.2.2 Effects on the epidermis
The main effects of surface-active agents on the epidermal
(stratum corneum) are:
-- delipidation of the skin surface or outer layer;
-- elution of natural moisturizing factor, which maintains the
water content of the outer layer;
-- denaturation of stratum corneum protein; and
-- increased permeability, swelling of the outer layer, and
inhibition of enzyme activities in the epidermis.
These effects and some others present a hazard to the skin; they
are described below.
In an investigation of the relationship between the irritating
potential of LAS in vivo and its ability to remove lipid from the
stratum corneum in vitro, LAS removed detectable levels of lipids
only at levels above the critical micelle concentration (0.04%). LAS
removed only small amounts of cholesterol, free fatty acids, the
esters of those materials, and possibly squalene. At concentrations
below that level, LAS can bind to and irritate the stratum corneum.
The clinical irritation produced by LAS is therefore unlikely to be
directly linked to extraction of lipid, and milder forms of
irritation may involve binding of LAS to and denaturation of keratin
as well as disruption of lipid (Froebe et al., 1990).
The results of the human arm immersion test with measurement of
eluted amino acids and protein, the skin permeation test, freeing of
sulfhydryl groups, and the patch test were compared for nine kinds
of surfactant, including LAS, ABS, AS, alcohol ethoxylate sulfate,
soap, nonionic surfactant, and amphoteric surfactant. LAS gave
intermediate reactions in the patch test and the permeation test and
showed a high level of sulfhydryl group freeing activity. The
results of the tests for evaluating surfactants did not agree with
those for the immersion test, which the author considered to provide
the best simulation of actual use (Polano, 1968).
In a number of studies, denaturation of outer layer proteins was
observed in vitro (Van Scott & Lyon, 1953; Harrold, 1959; Wood &
Bettley, 1971; Imokawa et al., 1974; Okamoto, 1974; Imokawa et al.,
1975b; Imokawa & Katsumi, 1976). Sodium dodecylbenzenesulfonate
stimulated penetration of sodium ions through isolated human
epidermis, partly because the detergent can denature proteins of the
epidermal stratum corneum (Wood & Bettley, 1971). Sodium laurate and
sodium lauryl sulfate were the most effective of several surfactants
in inducing swelling of the horny layer (Putterman et al., 1977).
The lysosome labilizing effects of surfactants, measured as the
release of enzyme from lysosomes, were shown to diminish in the
order cationic > anionic > nonionic surfactants (Imokawa &
Mishima, 1979). When ovalbumin was used as a simulated epidermis
protein, sodium lauryl sulfate was found to denature skin protein
extensively by exposing concealed sulfhydryl groups in LAS of alkyl
chain length C8-C16 (Blohm, 1957).
In immersion tests of the hand and the forearm up to 5 cm above
the wrist, falling off of skin scales diminished in the order:
sec-alkane sulfonate > LAS > AOS, alcohol ethoxylate sulfate
(Okamoto, 1974), but the distribution of carbon chain lengths among
the samples was not described. In a comparison of skin roughening by
a circulation method, the effects diminished in the order C12 AS
> C12 AOS > C12 sec-alkane sulfonate > C12 LAS (Imokawa
et al., 1974, 1975a,b). Skin roughening caused by several
surfactants that are components of commercial products was studied
by the method of Ito & Kakegawa (1972), in which various
concentrations are dripped onto the fingers. The effects diminished
in the order C10-C13 LAS = C12-C15 AS > C11, C13, C15
alcohol ethoxylate sulfate ( n = 0-3) > C14, C16, C18 AOS
> C11-C15 polyoxyethylene alkylether (Sadai et al., 1979).
A8.2.3 Hand eczema
The skin reaction to 0.04, 0.4, and 4.0% aqueous solutions of
LAS (10.0% C10, 34.3% C11, 31.5% C12, 24.7% C13) was tested
in a 24-h closed-patch test on the lower backs of 10 healthy
volunteers and 11 patients with hand eczema (progressive keratosis
palmaris). The incidence and intensity of skin reactions were
greater in the group with hand eczema, but the difference was not
statistically significant (Okamoto & Takase, 1976a,b).
In order to assess the possible etiological correlation between
exposure to LAS and hand eczema, 0.04, 0.4, and 4% aqueous solutions
of LAS were applied in 48-h closed-patch tests on the lower backs
of 20 women with hand eczema and 42 with other skin diseases. The
skin reaction was scored grossly from 0 to 5 on the basis of the
occurrence or intensity of erythema, papules, and vesicles. The
average score appeared to increase in parallel with the
concentration of LAS but did not differ between the groups with hand
eczema and other skin diseases (Sasagawa et al., 1978).
Nine proprietary household detergents were tested in 24-h
closed-patch tests on the lower backs of 160 women with hand eczema.
The surfactant concentrations in five of the products were: (i) 2%
ABS-Na, 15% LAS-Na; (ii) 2% ABS-Na, 14% LAS-Na; (iii) 17% LAS-Na,
12% alcohol ethoxylate sulfate; (iv) 11% ABS-Na, 11% LAS-Na; (v) 19%
LAS-Na. When the detergents were applied daily (for an unspecified
period) at an aqueous concentration of 0.175-0.8%, positive
responses were observed in 3.1% of the women, but they were
considered not to be allergic because the redness of the skin
disappeared completely within two days (Kawamura et al., 1970).
Three proprietary household detergents containing LAS were
tested in 24-h closed-patch tests on the forearms of 13 women with
'housewives' dermatitis' and 13 with other skin diseases. The
detergent was applied either undiluted or in a 0.2% aqueous
solution. Undiluted solutions of all three detergents caused mild to
moderate skin reactions, at incidences of 38.5, 48.1, and 73.1%,
which did not differ between the groups with housewives' dermatitis
and other skin diseases. The 0.2% aqueous solutions did not induce
skin reactions (Ishihara & Kinebuchi, 1967).
Two series of field tests were conducted to estimate if exposure
to a variety of synthetic detergent formulations was associated with
causation or aggravation of hand eczema in women. In the first
series, 162 female volunteers were divided into two groups and
instructed to wear a rubber glove on either the left or the right
hand while using the detergents. The test was conducted for one
month, and the gross appearance of hands before and after the test
period was compared. The relative intensity of noninflammatory
keratosis of the hands was increased in individuals in both groups
on hands that were covered and to a slightly greater extent on hands
that were uncovered. In the second series of tests, 881 housewives
were divided into three groups and instructed to use only one brand
of household detergent, containing LAS, AOS, or ABS during the test
period and to wear rubber gloves on both hands while using the
detergent. The test was conducted for 1.5 months, and the gross
appearance of hands before and after the test period was compared.
Skin roughness was not worsened in any of the three groups (Watanabe
et al., 1968).
A8.2.4 Occupational exposure
Sixty workers exposed at work to an atmosphere containing LAS at
8.64 mg/m3 were tested for serum lipid and sugar content and for
the activities of selected serum enzymes. The levels of total plasma
lipids and plasma cholesterol were slightly lower in the exposed
group than in controls, but no differences were noted for blood
sugar, plasma phospholipid, plasma lipoprotein, alpha-amylase,
leucine aminopeptidase, or pseudocholinesterase. The duration of
exposure before testing was not indicated (Rosner et al., 1973).
In an investigation of the asthmagenic properties of sodium
isononanoyl oxybenzene sulfonate, detergent industry workers were
also tested with LAS. Three workers previously exposed to sodium
isononanoyl oxybenzene sulfonate, three unexposed controls without
asthma, and three controls with asthma were challenged with
0.01-100 µg of LAS. No changes were seen after inhalation of LAS in
any of the subjects; but sodium isononanoyl oxybenzene sulfonate
induced asthmatic symptoms in the previously exposed workers and not
in the control groups (Stenton et al., 1990).
A8.2.5 Accidental or suicidal ingestion
No symptoms were seen in four cases of accidental ingestion of
unknown amounts of a household synthetic detergent containing LAS as
the main component (Hironaga, 1979).
A 32-year-old woman who had ingested 160 ml of a 21% aqueous
solution of LAS with suicidal intent showed transient, slight mental
confusion, vomiting, pharyngeal pain, hypotension, decreased plasma
cholinesterase activity, and increased urinary urobilinogen, but all
of these symptoms disappeared rapidly (Ichihara et al., 1967).
In a review of 1 581 540 cases of human exposure to a wide range
of chemicals reported by the United States Poison Control Centers in
1989, 7983 people had been exposed to household automatic dishwasher
preparations (alkali, anionic or nonionic, other or unknown) and 506
had required treatment in a health facility; 8950 had been exposed
to household cleansers, with 894 requiring treatment; 12 876 had
been exposed to laundry preparations, with 1542 treated; and 621 had
been exposed to industrial detergents (anionic, cationic, nonionic),
with 321 cases requiring treatment. There were no deaths, and only
12 of the treated cases were classified as 'major outcome'.
Virtually all the reports involved accidental exposure. The
compositions of the cleaning preparations, routes of exposure, and
clinical descriptions were not provided (Litovitz et al., 1990).
A9. EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND THE FIELD
Section summary
LAS have been tested extensively, both in the laboratory and
under field conditions, but the following aspects must be considered
in interpreting test results. Comparison of the results of tests
carried out on either mixtures of homologues of LAS or LAS of
specified chain length is restricted, because the toxicity of LAS is
influenced by the chain length, and homologues of lower chain length
are less toxic than those with longer chains; furthermore, chain
length was rarely specified in older studies. Studies of the effects
of formulations of LAS on environmental biota are not included in
this section.
Organisms are not exposed to a constant concentration of LAS in
water, owing to the high adsorptivity and biodegradability of LAS.
As LAS are adsorbed on suspended solids or food particles, they have
reduced bioavailability. The adsorption kinetics of LAS also depend
on the chain length of the homologues. Studies of aquatic toxicity
involving flow-through or static renewal (at least daily) should
therefore be given more prominence than studies based on static
conditions, although flow-through and static renewal cannot be used
in (semi-) chronic studies of lower organisms, such as daphnia.
Studies in which the actual concentration was measured should
likewise be given more consideration than those that rely on nominal
concentrations.
The effects of LAS on the aquatic environment have been studied
in short- and long-term studies in the laboratory and under more
realistic conditions: micro- and mesocosm and field studies. In
general, a decrease in alkyl chain length or a more internal
position of the phenyl group is accompanied by a decrease in
toxicity. Data on fish and daphnia indicate that a decrease in chain
length of one unit (e.g. C12 to C11) is accompanied by an
approximately 50% decrease in toxicity, but there is no linear
relationship between chain length and toxicity. In aquatic
microorganisms, the effects are strongly related to variables such
as the type of test system and use of mixed cultures as opposed to
individual species. EC50 values range from 0.5 mg/litre (single
species) to > 1000 mg/litre.
In freshwater fish, the acute LC50 values of C8-C15 LAS
are 0.1-125 mg/litre. The chronic L(E)C50 values of LAS (C11.7 and
not specified) in two species tested were 2.4 and 11 mg/litre, and
NOECs ranging from 0.11 to 8.4 mg/litre have been reported for
C11.2-C13 (or not specified). Marine fish appear to be more
sensitive, with acute LC50 values for C11.7 (or not specified)
in six species of 0.05-7 mg/litre, chronic LC50 values for LAS of
unspecified chain length in two species of 0.01-1 mg/litre, and an
NOEC for C12 in one species of < 0.02 mg/litre.
Results in aquatic plants are also species dependent. In
freshwater plants, the EC50 values for LAS (with chain lengths
shown in parentheses) were 10-235 mg/litre for green algae
(C10-C14), 5-56 mg/litre for blue algae (C11.1-C13),
1.4-50 mg/litre for diatoms (C11.6-C13), and 2.7-4.9 mg/litre
for macrophytes (C11.8). Marine algae appear to be even more
sensitive. There is probably no linear relationship between chain
length and toxicity to algae.
The effects of LAS on freshwater algae have also been tested
under realistic conditions in systems with various trophic levels,
comprising enclosures in lakes (lower organisms), model ecosystems
(sediment: water systems), a river below and above a wastewater
treatment plant outfall, and experimental streams. In general, C12
LAS were used. Algae were more sensitive in summer, when the 3-h
EC50 values with regard to photosynthesis were 0.2-8.1 mg/litre,
whereas studies of model ecosystems showed no effects on the
relative abundance of algal communities at 0.35 mg/litre. No effects
were seen in these studies at 0.24-5 mg/litre, depending on the
organism and parameter tested.
In aquatic invertebrates, the acute L(E)C50 values were
4.6-200 mg/litre for molluscs (either C13 or not specified),
0.12-27 mg/litre for crustaceans (C11.2-C18 or not specified),
1.7-16 mg/litre for worms (C11.8 or not specified), and
1.4-270 mg/litre for insects (C10-C15). The chronic L(E)C50
values were 2.2 mg/litre for insects (C11.8) and 1.1-2.3 mg/litre
for crustaceans (C11.8-C13). The chronic NOEC for crustaceans,
on the basis of lethality or reproduction, was 0.2-10 mg/litre
(C11.8 or not specified). Marine invertebrates are more sensitive,
with LC50 values of 1 to >100 mg/litre (almost all C12) and
NOEC values of 0.025-0.4 mg/litre (chain lengths not specified).
Biodegradation products and by-products of LAS are 10-100 times
less toxic than the parent compound.
Fewer data are available on the effects of LAS in the
terrestrial environment. For the plant species tested, the NOEC
values were < 10-20 mg/litre in nutrient solutions and 100 mg/kg
(C10-C13) for growth of plants in soils. The 14-day LC50 for
earthworms was > 1000 mg/kg.
One study in which chickens were treated in the diet resulted in
an NOEC based on egg quality of > 200 mg/kg.
A9.1 Effect of chain length on the toxicity of linear
alkylbenzene sulfonates
The ecotoxicity of homologues of LAS varies according to the
length of the alkyl chain and the position of the benzene ring on
this chain. In general, homologues with longer chains are more
ecotoxic than shorter ones, and ecotoxicity increases with the
proximity of the benzene ring to the end of the chain. The results
of studies on the effect of LAS chain length on acute toxicity to
fish are presented in Table 23.
The effect of chain length can also be seen on the basis of
quantitative structure-activity relationships (Roberts, 1989, 1991)
calculated from the octanol-water partition coefficients of
homologues of LAS. The slope of the relationship varied from 0.64 to
0.78; therefore, using an average slope of 0.70, it was calculated
that a decrease in chain length from C12 to C11 reduced the
aquatic toxicity of LAS by a factor of 2.4, with a corresponding
decrease in the octanol-water partition coefficient of 0.54.
Table 23. Effect of the chain length of linear alkylbenzene sulfonates (LAS)
on their acute toxicity to freshwater fish
Homologue Fathead minnow Goldfish Guppy Golden orfe
of LAS Pimephales Carassius Lebistes Idus idus
promelas auratus reticulatus melanotus
48-h LC50 6-h LC50 LC50 (mg/litre)c 96-h LC50
(mg/litre)a (mg/litre)b (mg/litre)d
C10 43.0 61.0 50 16.6
C11 16.0 22.5 6.5
C12 4.7 8.5 5 2.6
C13 0.4 3.3 0.57
C14 0.4 1 0.26
C16 0.087 1 0.68
C18 0.38 15
a From Kimerle & Swisher (1977)
b From Gafa (1974)
c From Borstlap (1967)
d From Hirsch (1963)
A9.2 Microorganisms
No adverse effects were seen on the performance of
laboratory-scale activated sludge units after addition of <
20 mg/litre LAS. At 50 mg/litre, nitrification was decreased in
extended aeration units that were treating synthetic sewage (Janicke
& Niemitz, 1973). A bacterium similar to Klebsiella pneumoniae
isolated from sewage degraded LAS at a concentration of 10 ml/litre,
but a concentration of 20 ml/litre inhibited the growth of the
bacterium by 39% (Hong et al., 1984).
The toxicity of microorganisms in activated sludge increases
with the length of the alkyl chain up to approximately C12 and
then decreases (Table 24), presumably because of decreased
bioavailability (e.g. greater sorption of these higher chain
lengths) (Verge et al., 1993).
Table 24. Results of tests for the inhibition of activated
sludge by the sodium salt of linear alkylbenzene
sulfonates (LAS)
LAS Chain length 3-h EC50 (mg/litre)
Pure homologues C10 1042-1200
C11 740-782
C12 500-723
C13 700-795
C14 900-1045
Commercial
formulations C11 760
C11.6 550
C13 650
From Verge et al. (1993)
A mixed bacterial culture was acclimatized to 10 mg/litre LAS
(C9-C14) and was then maintained in either river water, forest
soil, or wastewater from a detergent plant, the concentration of LAS
being increased every five days. At 20.8 and 46 mg/litre, no effect
was reported on the specific growth rate of the bacteria; however,
at 70 mg/litre, the growth rate was inhibited by 18%, and at
95 mg/litre growth was almost zero. Concentrations of 186 and
465 mg/litre LAS inhibited growth completely (Hrsak et al., 1981).
The acute toxicity of LAS (C9-C14) in naturally occurring
bacteria was studied in freshwater and seawater samples by measuring
3H-thymidine incorporation. The EC50 values were 0.5-1.66 mg/litre
for all samples. Toxicity was found to increase with an increasing
relative abundance of longer carbon chains (Martinez et al., 1989).
For bacteria collected from the Rhone River plume (an estuarine
area) and exposed to LAS, the EC50, based on 3H-thymidine
incorporation, was 11.9 mg/litre (Martinez et al., 1991).
The 8-h EC50, based on specific growth rate, of Pseudomonas
fluorescens in solutions of C11.1 LAS under static conditions
was 3200-5600 mg/litre (Canton & Slooff, 1982).
The effect of C11.6 LAS on the structure and function of
microbial communities was studied in a flow-through model ecosystem
containing several trophic levels at concentrations of 0.5 or
5 mg/litre. LAS had no effect on microbial structure at either dose
level, but at 5 mg/litre it inhibited the degradation of both
glucose and LAS. In an experiment in which LAS were supplied in
sewage, neither microbial structure nor function was affected
(Larson & Maki, 1982).
The effects of LAS on the microbial activity of soils were
studied on the basis of Fe[III] reduction. The no-effect-level was
found to be 250 mg/kg; the EC50 was about 500 mg/kg in a strongly
adsorbing soil and 33-55 mg/kg in a poorly adsorbing soil (Welp &
Brummer, 1985).
LAS at concentrations of 0.8-50 g/m2 had no effect on
respiration of loamy soil, sandy soil, or sandy soil irrigated with
wastewater for one or 14 days (Litz et al., 1987).
A9.3 Aquatic organisms
A9.3.1 Aquatic plants
A9.3.1.1 Freshwater algae and cyanobacteria
The 96-h EC50 values for C13 LAS on population growth were
116 mg/litre for the green alga Selenastrum capricornutum,
5 mg/litre for the blue-green alga Microcystis aeruginosa, and
1.4 mg/litre for the diatom Navicula pelliculosa. The EC50
values for C12 LAS were 29 mg/litre for Selenastrum and
0.9 mg/litre for Microcystis (Lewis & Hamm, 1986). The EC50 for
C11.7 LAS on growth of Selenastrum was reported to be 83
mg/litre (Konno & Wakabayashi, 1987). The EC50 values for C11.6
LAS were found to be 50-100 mg/litre for Selenastrum,
10-20 mg/litre for Mycrocystis, and 20-50 mg/litre for the diatom
Nitzschia fonticola (Yamane et al., 1984). The seven-day EC50
for C12 LAS in the green alga Chlorella pyrenoidosa, based on
growth, was 10 mg/litre (Kondo et al., 1983).
The 96-h EC50 values in algae grown in solutions of C11.1
LAS under static conditions, measured as biomass, were
32-56 mg/litre for Microcystis aeruginosa and 18-32 mg/litre for
Chlorella vulgaris (Canton & Slooff, 1982).
A study of the toxicity of various formulations of LAS to the
algae Scenedesmus subspicatus and Selenastrum capricornutum
(Table 25) indicated that commercial mixtures are as or slightly
less toxic than homologues. This finding may be due to a difference
in the sensitivity of the two algae, since those tested with the
homologues were of a different origin than those tested with
commercial LAS (Verge et al., 1993).
Table 25. Results of tests for the toxicity of the sodium salt
of linear alkylbenzene sulfonates (LAS) in algae
LAS Chain length 72-h EC50 (mg/litre)
Pure homologues C10 235
C11 118
C12 62
C13 33
C14 18
Commercial
formulations C11 80
C11.6 80
C13 62
From Verge et al. (1993)
LAS (chain length not specified) significantly reduced the
growth of the green alga Selenastrum capricornutum at a
concentration of 40 mg/litre or more. A significant decrease in
growth was also noted at 10 mg/litre, but no significant effect was
observed at 20 or 30 mg/litre (Nyberg, 1988).
A9.3.1.2 Marine algae
Growth of Gymnodinium breve was reduced by 69% rafter nine days'
exposure to C12 LAS (Kutt & Martin, 1977). These results were
confirmed in a study in which C13 LAS were introduced at the
bottom or surface of a water column: Exposure to LAS at
concentrations > 0.025 mg/litre inhibited growth completely within
two days (Hitchcock & Martin, 1977). These results suggest that
Gymnodinium breve is more sensitive to the effects of LAS than
other algae.
For C11.7 LAS, the seven-day EC50 for growth and the two-day
EC50 for ATP activity on the marine diatom Thalassiosira
pseudonana were both 10 mg/litre (Kondo et al., 1983).
Exposure of the alga Porhyra yezoensis, a standard test
species in Japan, to LAS (C10-C14) under semi-static conditions
gave a 10-day E50 (based on growth) of 0.56 mg/litre (Takita,
1985).
A9.3.1.3 Macrophytes
The seven-day EC50 values for C11.8 LAS on the duckweed
Lemna minor under flow-through conditions were 2.7 mg/litre for
frond count, 3.6 mg/litre for dry weight, and 4.9 mg/litre for root
length. The time-independent EC50 for growth rate and doubling
time was 4.8 mg/litre (Bishop & Perry, 1981).
A9.3.2 Aquatic invertebrates
A9.3.2.1 Acute toxicity
The acute toxicity of LAS to aquatic invertebrates is summarized
in Tables 26 and 27. For marine invertebrates, the 96-h LC50
values for C12 LAS range from 3 mg/litre for barnacles to >
100 mg/litre for several other species (Table 26). Freshwater
invertebrates show a range of 48-h LC50 values from 0.11 mg/litre
(C16) for a daphnid to 270 mg/litre (C11.8) for an isopod (Table
27). Several marine invertebrate species are more sensitivite to LAS
at the larval stage than as adults (Table 26).
Freshwater mussels (Anodonta cygnea) were more sensitive to
LAS during the reproductive period than during the non-reproductive
period, the 96-h LC50 being reduced from 200 to 50 mg/litre
(Bressan et al., 1989).
Studies with Daphnia magna revealed a correlation between
chain length and toxicity. The acute toxicity (24-h and 48-h LC50)
of LAS to Daphnia magna increased with chain length between C10
and C14 (Kimerle & Swisher, 1977) and with chain lengths between
C10 and C16 (Maki & Bishop, 1979), although similar values were
obtained for C16 and C18 homologues. No significant difference
in sensitivity was seen between Daphnia magna and Daphnia pulex.
A similar result was obtained with homologue mixtures (Martinez et
al., 1989): toxicity was correlated with the homologues in which
long chains were the most abundant.
Partial biodegradation of LAS significantly reduces the specific
toxicity (by unit weight) of the remaining LAS to Daphnia magna.
For example, LAS with a high relative molecular mass and a 48-h
LC50 of 2 mg/litre had an LC50 of 30-40 mg/litre after 80-85%
degradation (Kimerle & Swisher, 1977); the longer homologues and
more terminal isomers, which are the most toxic, are therefore also
the more readily biodegraded. Shorter carboxylates formed during the
degradation of LAS were three to four orders of magnitude less toxic
than LAS (Swisher et al., 1978). Other workers also found a
Table 26. Acute toxicity of linear alkylbenzene sulfonates (LAS) to estuarine and marine invertebrates
Organism Size or Static or Temp. Salinity LAS chain End-point Concentration Reference
age flow (°C) (%) length (mg/litre)a
Sea squirt Larva Static 20 NS 6-h LC50 1 Renzoni (1974)
(Ciona intestinalis)
Common mussel Static 6-8 32-34 C12 96-h LC50 > 100 Swedmark et
(Mytilus edulis) Static 15-17 32-34 C12 96-h LC50 50 al. (1971)
Mussel Staticr 18 35 NS 48-h LC50 39.8 Bressan et al.
(Mytilus galloprovincialis) Adult 18 35 NS 96-h LC50 1.66 (1989)
Cockle Static 6-8 32-34 C12 96-h LC50 5 Swedmark et
(Cardium edule) Juvenile Static 15-17 32-34 C12 96-h LC50 5 al. (1971)
Clam Static 6-8 32-34 C12 96-h LC50 70
(Mya arenaria) Static 15-17 32-34 C12 96-h LC50 < 25
Scallop Static 6-8 32-34 C12 96-h LC50 < 5
(Pecten maximus)
Scallop Static 15-17 32-34 C12 96-h LC50 < 5
Decapod Static 15-17 32-34 C12 96-h LC50 50
(Leander adspersus) Intermoult Static 6-8 32-34 C12 96-h LC50 50
Postmoult Static 6-8 32-34 C12 96-h LC50 25
Hermit crab Static 6-8 32-34 C12 96-h LC50 > 100
(Eupagurus bernhardus)
Table 26 (contd)
Organism Size or Static or Temp. Salinity LAS chain End-point Concentration Reference
age flow (°C) (%) length (mg/litre)a
Spider crab Larva Static 6-8 32-34 C12 96-h LC50 9
(Hyas araneus) Adult Static 6-8 32-34 C12 96-h LC50 > 100
Shore crab Static 6-8 32-34 C12 96-h LC50 > 100
(Carcinus maenus)
Barnacle Larva Static 6-8 32-34 C12 96-h LC50 3
(Balanus balanoides) Adult Static 6-8 32-34 C12 96-h LC50 50
Brine shrimp Static 25 C11-C13 24-h LC50 33 Price et al.
(Artemia salina) (1974)
Static: water unchanged for duration of test; NS, not specified; staticr, static renewal: water changed every 12 h; flow, flow-through
conditions: LAS concentration in water maintained continuously
a Based on nominal concentration
Table 27. Acute toxicity of linear alkylbenzene sulfonates (LAS) to freshwater invertebrates
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Bivalve mollusc 11 cm Staticr 18 8.0 NS 96-h LC50 200b Bressan et al.
(Anodonta cygnea) 18 8.0 NS 96-h LC50 50b,c (1989)
Bivalve mollusc 9 cm Staticr 18 8.0 NS 96-h LC50 182.5b
(Unio elongatulus)
Snail Static 21 62 7.3 av. C13 24-h LC50 4.6b Dolan &
(Gonobasis sp.) Hendricks (1976)
Snail (Physa integra) Flow 15 41-47 7.5-7.7 NS 96-h LC50 9b Arthur (1970)
Amphipod (Gammarus Flow 15 41-47 7.5-7.7 NS 96-h LC50 7b
pseudolimnaeus)
Amphipod 4.3 mm Static 22 165 7.9-8.4 C11.8 48-h LC50 3.3b Lewis &
(Gammarus sp.) Suprenant (1983)
Campeloma decisum Flow 15 41-47 7.5-7.7 NS 96-h LC50 27b Arthur (1970)
Water flea < 24 h Static 20 25 C11.7 24-h LC50 17 Wakabayashi
(Daphnia magna) et al. (1988)
< 24 h Static 21 120 7.4 C10 48-h LC50 9.55d Maki &
Bishop (1979)
Table 27 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Water flea (contd) < 24 h Static 21 120 7.4 C11 48-h LC50 1.15d
(Daphnia magna) < 24 h Static 21 120 7.4 C12 48-h LC50 5.88-6.84d
< 24 h Static 21 120 7.4 C13 48-h LC50 2.63d
< 24 h Static 21 120 7.4 C14 48-h LC50 0.68-0.8d
< 24 h Static 21 120 7.4 C16 48-h LC50 0.11-0.2d
< 24 h Static 21 120 7.4 C18 48-h LC50 0.12d
< 18 h Static C13.3 48-h LC50 2.3b Kimerle &
< 18 h Static C10 48-h LC50 12.3b Swisher (1977)
< 18 h Static C11 48-h LC50 5.7b
< 18 h Static C12 48-h LC50 3.5b
< 18 h Static C13 48-h LC50 2.0b
< 18 h Static C14 48-h LC50 0.7b
< 24 h Static 19 C11.2 48-h LC50 18-32b Canton &
Slooff (1982)
< 24 h Static 21 131 7.4-7.8 C11.8 48-h LC50 4.8d Lewis (1983)
< 24 h Static 22 165 7.9-8.4 C11.8 48-h LC50 1.8-5.6b Lewis &
Suprenant
(1983)
< 24 h Static 21 295-310 7.3-8.4 C11.8 48-h LC50 3.6-4.7b Taylor (1985)
< 48 h Static 22 241 7.8 C11 48-h EC50 2.2b,e Barera &
Adams (1983)
Flow C11.8 48-h LC50 4.4d Bishop &
Perry (1981)
< 12 h Flow 21 120 7.4 C11.8 96-h LC50 23.94d Maki (1979a)
< 12 h Flow 21 120 7.4 C13 48-h LC50 2.19d
Table 27 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Water flea < 24 h Static 20 25 C11.7 24-h LC50 18 Wakabayashi
(Daphnia pulex) et al. (1988)
< 24 h Static 21 120 7.4 C12 48-h LC50 8.62d Maki &
Bishop (1979)
< 24 h Static 21 120 7.4 C14 48-h LC50 0.59d
< 24 h Static 21 120 7.4 C16 48-h LC50 0.15d
Oligochaete (Dero sp.) 6.0 mm Static 22 165 7.9-8.4 C11.8 48-h LC50 1.7b Lewis &
Suprenant
(1983)
Roundworm (nematode) 0.3 mm Static 22 165 7.9-8.4 C11.8 48-h LC50 1.7b
(Rhabditis sp.)
Flatworm 3.4 mm Static 22 165 7.9-8.4 C11.8 48-h LC50 1.8b
(Dugesia sp.)
Branchiura sowerbyi Staticr 10 25 8.0 NS 96-h LC50 10.8b,f Bressan et al.
10 25 8.0 NS 96-h LC50 4.4b (1989)
Worm (Limnodrilus Staticr 10 25 8.0 NS 96-h LC50 7.8b,f
hoffmeisteri) 10 25 8.0 NS 96-h LC50 2.0b
Isopod (Asellus sp.) 5.3 mm Static 22 165 7.9-8.4 C11.8 48-h LC50 1.8b Lewis &
Suprenant
(1983)
Table 27 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Midge (Chironomus Larva Flow 22 150 7.8-8.4 C11.8 72-h LC50 2.2d Pittinger
riparius) et al. (1989)
Midge (Paratanytarsus 3.6 mm Static 2 165 7.9-8.4 C11.8 48-h LC50 1.8d Lewis &
parthenogenica) Suprenant
(1983)
Mosquito (Aedes Larva Static C10-13 24-h LC50 6b Van Emden et
aegypti) Larva Static C10-15 24-h LC50 2b al. (1974)
3-4 d Static 23 C11.1 48-h LC50 56-100b Canton &
Slooff (1982)
Mayfly Larva Static 10 53 7.5-7.8 C11.6 24-h LC50 13.6b Dolan et al.
(Isonychia sp.) Larva Static 10 53 7.5-7.8 C11.6 48-h LC50 10.4b (1974)
Larva Static 10 53 7.5-7.8 C11.6 96-h LC50 5.33b
Larva Static 10 53 7.5-7.8 C13.1 24-h LC50 4.19b
Larva Static 10 53 7.5-7.8 C11.6 48-h LC50 12.47b
Larva Static 10 53 7.5-7.8 C11.6 96-h LC50 1.36b
Staticr, static renewal: water changed every 12 h; NS, not specified; flow, flow-through conditions: LAS concentration in water maintained
continuously; static: water unchanged for duration of test
a mg/litre CaCO3
b Based on nominal concentration
c Test performed during the reproductive period
d Based on measured concentrations
e Based on immobilization
f Organism exposed in the presence of sediment
reduction in the acute toxicity of LAS to Daphnia magna during
primary degradation (Gard-Terech & Palla, 1986).
Increasing hardness also increased the acute toxicity (48-h
LC50) of C11.8 LAS from a nominal concentration of 7.1 mg/litre
at 25 mg/litre CaCO3 to 4.0 mg/litre at 350 mg/litre CaCO3;
however, significant additional physiological stress was induced if
the hardness of the culture water was significantly different from
that of the test water. Pre-exposure to 0.4 mg/litre LAS (one-tenth
of the 48-h LC50) for up to seven generations (14 weeks) had no
significant effect on the susceptibility of daphnids to acute
exposures (Maki & Bishop, 1979).
Loading density, ranging from 10 daphnids per 20 ml to 20
daphnids per 1000 ml, had no significant effect on the acute
toxicity of C11.8 LAS for Daphnia magna (Lewis, 1983). Daphnids
fed a diet containing Selenastrum had a significant, twofold
decrease in acute toxicity due to C11.8 LAS in comparison with
unfed daphnids (Taylor, 1985). The presence of sediment reduced the
acute toxicity of LAS to the oligochaete worms Branchiura sowerbyi
and Limnodrilus hoffmeisteri. The NOEC and LOEC for B. sowerbyi
were 2.5 times higher in the presence of sediment, and those for
L. hoffmeisteri were 4-4.5 times higher (Bressan et al., 1989; see
also Table 27).
The 96-h EC50 values for duplicate studies of the effect of
LAS on attachment of the podia of the sea urchin Hemicentrotus
pulcherrimus were 3.7 and 3.8 mg/litre (Lee & Park, 1984).
The data from other studies (Lal et al., 1983, 1984a,b; Misra et
al., 1984; Chattopadhyay & Konar, 1985; Misra et al., 1985; Devi &
Devi, 1986; Misra et al., 1987, 1989a,b, 1991) could not be
adequately interpreted because of deficiencies in the data or
method, including inadequate characterization of the test material
with regard to chain-length distribution and use of test material in
an acidified form. The range of values for toxicity reported in
these studies was 10-100 times greater than that in numerous studies
of the same or similar species, and the high values have not been
verified by these or other researchers. As the toxic effects
reported are not considered to be representative of those of
commercial LAS, the data were not used in evaluating the
environmental effects of LAS.
A 72-h LC50 of 2.2 mg/litre was reported for C11.8 LAS in
newly hatched larvae of the midge (Chironomus riparius) (Pittinger
et al., 1989).
A9.3.2.2 Short-term and long-term toxicity
The 21-day LC50 for the water flea (Daphnia magna) was
18 mg/litre, and the NOEC, based on survival, was 10 mg/litre under
static renewal conditions. The 21-day EC50, based on reproduction,
was estimated to be > 10 mg/litre (Canton & Slooff, 1982). The
14-day EC50 for C12 LAS in Daphnia carinata, based on
reproduction, was 16.8 mg/litre (Hattori et al., 1984).
Diet had a significant effect on the sensitivity of Daphnia
magna to the chronic toxicity of C11.8 LAS. The NOEC values
showed a threefold variation of 1.2-3.2 mg/litre and the 21-day
LC50 values a twofold variation of 2.2-4.7 mg/litre with diet. A
threefold variation in toxicity in tests in Daphnia is not,
however, unusual (Taylor, 1985).
Under continuous-flow conditions, a 21-day LC50 value of
1.67 mg/litre was found for daphnids (Daphnia magna) exposed to
C11.8 LAS and 1.17 mg/litre for those exposed to C13 LAS. The
EC50 values for reproductive toxicity were 1.5 mg/litre for
C11.8 LAS and 11.1 mg/litre for C13 with respect to total young
production, 2.3 mg/litre for C11.8 and 1.4.1 mg/litre for C13
for average brood size, and 2.31 mg/litre for C11.8 and
1.29 mg/litre for C13 for percentage of days on which reproduction
occurred (Maki, 1979a).
Campeloma decisum, Gammarus pseudolimnaeus, and Physa integra
were exposed to LAS at concentrations of 0.2-4.4 mg/litre for six
weeks; amphipods were exposed for a further 15 weeks. Survival,
growth, reproduction, feeding, and mobility were studied. The
maximum acceptable concentrations of LAS were found to be
0.2-0.4 mg/litre for Gammarus and 0.4-1.0 mg/litre for Campeloma;
P. integra were not significantly affected (Arthur, 1970).
Fertilized eggs of sea urchins (Paracentrotus lividus) were
treated with LAS at concentrations of 0-0.5 mg/litre for 40 days.
The pattern of embryonic development was unaffected, but the mean
length of the somatic rods of the echinoplutei were reduced
successively with increasing LAS concentrations. A significant
reduction in growth occurred at doses between 0.35 and 0.4 mg/litre;
above 0.45 mg/litre, alterations in skeletal development were
induced (Bressan et al., 1989).
Oligochaete worms (B. sowerbyi) were maintained in LAS at a
concentration of 0.5, 2.5, or 5.0 mg/litre for up to 140 days in the
presence of sediment. Exposed worms laid fewer cocoons and eggs, but
the worms exposed to 5 mg/litre were the least affected. The
percentage of degenerated cocoons, the percentage of worms hatching,
the mean number of eggs per cocoon, and the mean embryonic
development time were all unaffected by treatment. Worms exposed via
the sediment only were not affected (Bressan et al., 1989).
Growth of mussels (Mytilus galloprovincialis) exposed to LAS
at a concentration of 0.25 or 0.5 mg/litre for 220 days, expressed
as mean length of the major axis of the shell, was significantly
slowed ( p < 0.001). The mean (± SE) increments in growth were:
control, 3.11 ± 0.34; 0.25 mg/litre, 1.71 ± 0.15; 0.5 mg/litre,
1.48 ± 0.16 (Bressan et al., 1989).
Eggs of the common mussel, M. edulis, were exposed from the
time of fertilization for 240 h. Fertility was decreased at the
lowest concentration of 0.05 mg/litre and fertilization did not take
place at concentrations in excess of 1 mg/litre. LAS at
concentrations > 0.3 mg/litre inhibited the development of mussel
larvae by delaying the transitory stages of larval development.
Reduced growth rates were observed at concentrations > 0.1 mg/litre
(Granmo, 1972).
Newly fertilized eggs of American oysters (Crassostrea
virginica) were exposed to LAS (chain length not specified, but
likely to be C13) for 48 h. The percentage of eggs that developed
normally was significantly reduced at concentrations greater than
0.025 mg/litre. The percentage survival of oyster larvae hatched in
'clean' water and exposed to LAS at a concentration of 1 mg/litre
for 10 days was significantly decreased, and growth (mean length)
was significantly reduced at 0.5 mg/litre (Calabrese & Davis, 1967).
Embryos of sea urchins (P. lividus) were exposed to LAS at
concentrations of 0.25-0.5 mg/litre from the time of fertilization
for 40 h. At concentrations > 0.45 mg/litre, skeletal development
was totally inhibited; a significant decrease was observed at
0.3 mg/litre. The effect of LAS was found to be maximal at the end
of gastrulation when calcium uptake is high (Bressan et al., 1991).
The effects of LAS were studied on the eggs and sperm of the sea
squirt Ciona intestinalis. Fertility and hatchability were
markedly reduced at 0.1 mg/litre when eggs and sperm were exposed
for the entire developmental period; however, if they were exposed
only before fertilization, fertility and hatchability were slightly
reduced at 0.1 mg/litre but markedly at 1 mg/litre. Male gametes
appeared to be particularly sensitive to the toxic effects of LAS
(Renzoni, 1974).
Two marine benthic filter feeders, the sea squirts Botryllus
schlosseri and Botrylloides leachi were exposed at different
periods of development to LAS. When larvae were exposed from
spawning for 6 h, the incidence of abnormal metamorphosis was
significantly increased at 1 mg/litre LAS for Botryllus and 2
mg/litre for Botrylloides. The frequency of spontaneously settled
larvae of both species also increased with exposure to LAS and
seemed to be a selective effect of LAS. The frequency was
significantly different from controls at 1 and 3 mg/litre for the
two species, respectively. In a second experiment, young colonies
were exposed to LAS for 15 days immediately after discharge by the
parental colony. Growth rates were significantly decreased at
0.5 mg/litre for Botryllus and at 0.25 mg/litre for Botrylloides.
When colonies were exposed from the end of metamorphosis, their
growth rates were similarly affected, but the mortality rate was
significantly lower. The effects of LAS thus appear to be exerted
mainly on the pelagic phase of the life cycle (Marin et al., 1991).
No significant reduction in egg hatching of midges (Chironomus
riparius) was seen at the highest concentration of C11.8 LAS
tested (18.9 mg/litre), but newly hatched larvae were more
sensitive, with a 72-h LC50 of 2.2 mg/litre. In bioassays of part
of the life cycle in a sediment and water system, the percentages of
winged adults emerging were monitored after continuous exposure of
larvae and pupae. The NOEC for sediment containing LAS was 319 mg/kg
(dry weight). In the absence of sediment, the NOEC was
2.40 mg/litre. Both tests were conducted for about 20 days
(Pittinger et al., 1989).
A9.3.2.3 Biochemical and physiological effects
Juvenile mussels (M. galloprovincialis) were exposed to LAS at
a concentration of 0.25 or 0.5 mg/litre for 220 days. Oxygen uptake
and the retention rate of neutral red (a measure of filtration rate)
were significantly decreased, but no effect was detected on nitrogen
excretion (measured as ammonia). When the experiment was repeated
over a seven-day period at a concentration of LAS of 1 or
1.5 mg/litre, no significant effect was seen on nitrogen metabolism
and the results for oxygen uptake were inconclusive. The filtration
rate was again significantly reduced when compared with that in
control mussels (Bressan et al., 1989).
The 48-h LC50 for lugworms (Arenicola marina) exposed to LAS
was calculated to be 12.5 mg/litre (95% confidence interval,
8.6-18.2). When tissues from a lugworm exposed to a concentration
close to that of the LC50 were examined for changes in morphology
by both light and electron microscopy, serious damage was reported
in the caudal epidermis, epidermic receptors, and gills; no effect
was reported in the thoracic epidermis or the intestine. In the
caudal epidermis, LAS destroyed the papillae, disrupting the
internal structure, occasionally displacing the musculature below
the papillae and thus giving it direct contact with seawater.
Deciliation of the epidermic receptors was also reported. These
effects were considered to indicate that the physiological response
of damaged epidermic receptors was reduced or blocked by exposure to
LAS. Changes in the morphology of the gills included destruction of
the epithelium and blood vessels, causing complete solubilization of
branch apexes, and development of holes at the base of the gills
(Conti, 1987).
A9.3.3 Fish
A9.3.3.1 Acute toxicity
The acute toxicity of LAS to fish is summarized in Tables 28 and
29. Only a few studies were available on marine fish, providing two
96-h LC50 values, 1 and 1.5 mg/litre LAS. Tests in various species
of freshwater fish gave a wide range of LC50 values: the 48-h
values ranged from 0.2 mg/litre for brown trout (Salmo trutta) to
125 mg/litre for the golden orfe (Idus idus memanotus), and the
96-h values ranged from 0.1 mg/litre for brown trout to 23 mg/litre
for white tilapia (Tilapia melanopleura).
The acute toxicity tended to increase with increasing carbon
chain length. Thus, C14 LAS were more acutely toxic to bluegill
(Lepomis macrochirus) than C12 compounds (Swisher et al., 1964);
the acute toxicity of LAS to the golden orfe increased with chain
length from C8 to C15 but decreased at C16 (Hirsch, 1963).;
and a similar trend was found for fathead minnows (Pimephalus
promelas) exposed to LAS with chain lengths of C10 to C14
(Kimerle & Swisher, 1977).
The 96-h LC50 values in bluegill (Lepomis macrochirus) were
0.64 mg/litre for C14 and 3 mg/litre for C12 LAS but
75 mg/litre for the intermediate degradation product,
sulfophenylundecanoic acid disodium salt (Swisher et al., 1964).
Biodegradation of LAS with a high relative molecular mass
progressively shifted the homologue distribution in favour of
shorter chain lengths and reduced the acute toxicity of the compound
to bluegill (Dolan & Hendricks, 1976). Similar findings were
reported for fathead minnow (Swisher et al., 1978), goldfish
(Carassius auratus) (Divo & Cardini, 1980) and zebra fish
(Brachydanio rerio) (Gard-Terech & Palla, 1986).
In rainbow trout (Oncorhynchus mykissi), addition of LAS
(C10-C15) to activated sludge plant effluent increased the
nominal 96-h LC50 from 0.36 to 29.5 mg/litre (Brown et al., 1978).
No deaths were observed among bluegill exposed for 4-11 days to
effluent from continuous-flow activated sludge units fed
100 mg/litre LAS (Swisher et al., 1964).
Water hardness was found to be the most significant
environmental factor in the acute toxicity of LAS to bluegill,
increasing with the level of hardness. At a water hardness of
15 mg/litre CaCO3, the mean LC50 was 4.25 mg/litre; at
290 mg/litre CaCO3, the LC50 was reduced to 2.85 mg/litre
(Hokanson & Smith, 1971). Similarly, when water hardness was
increased from 0 to 500 mg/litre CaCO3, the LC50 for C18 LAS
in goldfish was reduced from 15 to 5.7 mg/litre (Gafa, 1974).
Exposure of the freshwater bleeker (Puntius gonionotus) to LAS
gave 96-h LC50 values of 13.6 mg/litre at a water hardness of
50 mg/litre CaCO3, 11.8 mg/litre at 110 mg/litre CaCO3, and
11.4 mg/litre at 260 mg/litre CaCO3 (Eyanoer et al., 1985).
The toxicity of C11.7 LAS to the medaka (Oryzias latipes)
increased with increasing salinity, but the effect was less
pronounced than that of water hardness (Wakabayashi & Onizuka,
1986).
Temperature was reported to have no significant effect on the
acute toxicity of LAS (Hokanson & Smith, 1971), but in another study
increasing the water temperature from 28 to 35°C marginally
decreased the 96-h LC50 for the bleeker, from 11.8 to
11.5 mg/litre (Eyanoer et al., 1985).
A reduction in the dissolved oxygen concentration from 7.5 to
1.9 mg/litre reduced the 24-h LC50 in bluegill from 2.2 to
0.2 mg/litre. When the fish were first acclimatized to reduced
oxygen levels, the effect was less pronounced (Hokanson & Smith,
1971).
No significant effect on the acute toxicity of LAS to bluegill
was observed after a bentonite suspension was added to water at
concentrations of 0, 50, or 200 mg/litre (Hokanson & Smith, 1971).
Addition of gluten, however, reduced the 24-h and 48-h acute
toxicity of LAS to both himedaka (Oryzias latipes) and cobalt
suzume (Chrysiptera hollisi) (Iimori & Takita, 1979).
A9.3.3.2 Chronic toxicity
Exposure of the eggs of fathead minnows (Pimephales promelas) to
LAS from laying until all surviving eggs had hatched under
flow-through conditions gave a nine-day LC50 of 2.4 mg/litre,
which would result in an LC50 of 3.4 mg/litre after 24 h of
exposure (Pickering, 1966).
Eggs of cod (Gadus morhua) were exposed to LAS at a
concentration of 0.005, 0.02, 0.05, or 0.1 mg/litre from
fertilization until hatching under flow-through conditions. There
were no significant effects at 0.005 mg/litre. At a concentration of
0.02 mg/litre, only 42% of the embryos completely developed into
larvae, and there was an increased occurrence of tail malformations
in comparison with controls. At 0.05 mg/litre, few eggs developed to
embryos. No eggs developed to the blastula stage at a concentration
of 0.1 mg/litre. In a repetition of the test at 0.05 mg/litre, fewer
eggs and larvae died, but there was an increased frequency of
abnormal embryos and inactive and crippled larvae (Swedmark &
Granmo, 1981).
Eggs, larvae, and immature adult fathead minnows (Pimephales
promelas) were exposed to LAS at a concentration of 0.34, 0.63,
1.2, or 2.7 mg/litre for up to four months. No significant effect
was observed on the number of spawnings, the total number of eggs
produced, the mean number of spawnings per female, the mean number
of eggs per spawning, or the percentage hatchability; however, the
two highest concentrations significantly reduced the survival of fry
(Pickering & Thatcher, 1970).
The effects of C11.8 and C13 LAS on the number of females,
the number of spawnings, total number of eggs produced, and number
of eggs per female were also studied in the fathead minnow over a
period of one year. As C11.8 LAS had no significant effect on
these parameters at a concentration of 1.09 mg/litre and a water
hardness of 120 mg/litre CaCO3, the NOEC was 0.9 mg/litre; C13
LAS were more toxic, and the NOEC was 0.15 mg/litre. At a lower
water hardness (39 mg/litre), however, survival of larvae was
impaired at 0.74 mg/litre (Maki, 1979a). NOECs in the fathead minnow
in life-cycle and embryo-larval tests were dependent on mean alkyl
chain length: 5.1-8.4 mg/litre for C11.2, 0.48 mg/litre for
C11.7, and 0.11-0.25 mg/litre for C13.3 (Holman & Macek, 1980).
The LC50 value of LAS in the eggs of carp (Cyprinus carpio)
exposed from spawning to hatching was 11 mg/litre. In determinations
of the sensitivity of eggs at different stages of development after
spawning, the 24-h LC50 values were 15 mg/litre for eggs exposed
between 2 and 26 h, 25 mg/litre for exposure between 26 and 50 h,
and 32 mg/litre for exposure between 50 h and hatching (Kikuchi et
al., 1976).
Bluegill (Lepomis macrochirus) were exposed to LAS from
fertilization to yolk-sac resorption at a concentration of 1.8, 3.5,
4.6, or 5.5 mg/litre. The lowest concentrations did not affect
hatchability or survival. Survival among those exposed to
3.5 mg/litre which hatched successfully was significantly reduced
within two days of hatching, and 95% were dead by the end of the
experiment. Eggs exposed to 4.6 or 5.5 mg/litre failed to hatch
(Hokanson & Smith, 1971).
The NOEC of LAS in guppies (Poecilia reticulata), based on
mortality, behaviour, and growth over 28 days, was 3.2 mg/litre
(Canton & Slooff, 1982).
Studies of the short- and long-term toxicity of LAS to
freshwater and marine fish are summarized in Tables 28 and 29.
A9.3.3.3 Biochemical and physiological effects
The main injury to the gills of catfish (Heteropneustes
fossilis) exposed to LAS at 1 or 2.5 mg/litre was progressive
separation of the lamellae from their vascular components.
Table 28. Toxicity of linear alkylbenzene sulfonates (LAS) to freshwater fish
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Brown trout Flow 15 26-30 NS 48-h LC50 5.3 Reiff et al.
(Salmo trutta) Flow 15 26-30 NS 96-h LC50 4.6 (1979)
Flow 15 26-30 NS 48-h LC50 2.3
Flow 15 26-30 NS 96-h LC50 1.4
Flow 15 26-30 NS 48-h LC50 0.4
Flow 15 26-30 NS 96-h LC50 0.4
Flow 15 250 NS 48-h LC50 2
Flow 15 250 NS 96-h LC50 0.9
Flow 15 250 NS 48-h LC50 0.7-0.9
Flow 15 250 C10-C15 48-h LC50 0.2
Flow 15 250 C10-C15 96-h LC50 0.1
Masu trout 2 mo Staticr 8.5-9.6 30 C11.7 96-h LC50 4.4 Wakabayashi et
(Oncorhynchus masou) al. (1984)
Rainbow trout Flow 15 250 C12.6 96-h LC50 0.36b Brown et al.
(Oncorhynchus mykiss) (1978)
40 d Staticr 8.8-10.9 25 C11.7 96-h LC50 4.7 Wakabayashi et
al. (1984)
4 d Staticr 10 25 C11.7 96-h LC50 2.1 Wakabayashi &
19 d Staticr 10 25 C11.7 96-h LC50 3.4 Onizuka (1986)
Goldfish Static 20 C16 6-h LC50 61 Gafa (1974)
(Carassius auratus) Static 20 C17 6-h LC50 22.5
Static 20 C18 6-h LC50 8.5
Table 28 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Goldfish (contd) Static 20 C19 6-h LC50 3.3
(Carassius auratus) Static 20 C16-C19 6-h LC50 7.6
Static 20 C16-C19 6-h LC50 10
Static 20 C16-C19 6-h LC50 12.2
Static 20 100 NS 6-h LC50 8.2 Reiff et al.
Static 20 100 NS 6-h LC50 7 (1979)
Static 20 100 NS 6-h LC50 4.3
3.1-6.0 Flow 20-23 45-96 7.1-9.3 24-h LC50 7.6 Tsai & McKee
cm Flow 20-23 45-96 7.1-9.3 48-h LC50 7.5 (1978)
Flow 20-23 45-96 7.1-9.3 72-h LC50 7.0
Flow 20-23 45-96 7.1-9.3 96-h LC50 6.2
Bluegill sunfish 1.6 g Static 23 76 7.5 av. C13 48-h LC50 0.72b Dolan &
(Lepomis macrochirus) 1.6 g Static 23 76 7.5 av. C13 96-h LC50 0.72b Hendricks
(1976)
Flow 23 50 7.5 av. C13 96-h LC50 4c Thatcher &
Santner (1967)
Finger Static 25 15 C11.2 48-h LC50 4.0-4.5b Hokanson &
Finger Static 25 290 C11.2 48-h LC50 2.8-2.9b Smith (1971)
Flow C11.8 96-h LC50 1.7c Bishop & Perry
(1981)
Fathead minnow Static C13.3 48-h LC50 1.7b Kimerle &
(Pimephales promelas) Static C10 48-h LC50 43b Swisher (1977)
Static C11 48-h LC50 16b
Table 28 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Fathead minnow (contd) Static C12 48-h LC50 4.7b
Static C14 48-h LC50 0.4b
2-3 mo Static 40 C11.2 96-h LC50 12.3c Holman & Macek
2-3 mo Static 40 C11.7 96-h LC50 4.1c (1980)
2-3 mo Static 40 C13.3 96-h LC50 0.86
Static 25 48-h LC50 4.6 Pickering &
Static 25 96-h LC50 5.0 Thatcher (1970)
Flow 15 43 7.2-7.9 96-h LC50 3.4 McKim et al.
(1975)
Flow 23 50 7.5 96-h LC50 4.2 Thatcher &
Santner (1967)
Flow 25 96-h LC50 4.2-4.5 Pickering &
Thatcher (1970)
2.5 cm Flow 18 116 7.9 C12 96-h LC50 3.5 Solon et al.
(1969)
Harlequin fish Flow 20 20 NS 48-h LC50 7.6 Reiff et al.
(Rasbora heteromorpha) Flow 20 20 NS 96-h LC50 6.1 (1979)
Flow 20 20 NS 48-h LC50 5.1
Flow 20 20 NS 96-h LC50 4.6
Flow 20 20 C10-C15 48-h LC50 0.9
Flow 20 20 NS 96-h LC50 0.7
Carp (Cyprinus carpio) 4.4 mg Static 22 25 7 C11.7 12-h LC50 5.6 Kikuchi et al.
Static 22 25 7 C11.7 48-h LC50 5.6 (1976)
Table 28 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Carp (contd) 3.5-5.5 Static 21 7.5-7.8 NS 48-h LC50 6.8 Lopez-Zavala
cm Static 21 7.5-7.8 NS 96-h LC50 5.0 et al. (1975)
7 d Staticr 22 25 7.0 C11.7 48-h LC50 5.6 Arima et al.
6 mo Staticr 22 25 6.5-7.1 C11.7 48-h LC50 10 (1981)
50 d Staticr 21 75 C11.7 96-h LC50 4.4 Wakabayashi et
al. (1984)
2 d Staticr 20 25 C11.7 96-h LC50 4.6 Wakabayashi &
15 d Staticr 20 25 C11.7 96-h LC50 2.6 Onizuka (1986)
White tilapia 5-7 cm Static 21 7.5-7.8 NS 48-h LC50 26 Lopez-Zavala
(Tilapia melanopleura) 5-7 cm Static 21 7.5-7.8 NS 96-h LC50 23 et al. (1975)
Guppy 3-4 wk Staticr 23 C8-C14 96-h LC50 5.6-10 Canton & Slooff
(Poecilia reticulata) (1982)
Northern pike Flow 15 43 7.2-7.9 96-h LC50 3.7 McKim et al.
(Esox lucius) (1975)
White sucker Flow 15 43 7.2-7.9 96-h LC50 4 McKim et al.
(Catostomus commersoni) (1975)
Golden orfe Static 18-20 C8 48-h LC50 125 Hirsch (1963)
(Idus idus melanotus) Static 18-20 C9 48-h LC50 88
Static 18-20 C10 48-h LC50 16.6
Table 28 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Golden orfe (contd) Static 18-20 C11 48-h LC50 6.6
Static 18-20 C12 48-h LC50 2.6
Static 18-20 C13 48-h LC50 0.57
Static 18-20 C15 48-h LC50 0.23
Static 18-20 C16 48-h LC50 0.68
1.2-1.8 g Static 20 NS 48-h LC50 3.94 Mann (1976)
Static 20 NS 48-h LC50 1.85
Static 20 NS 48-h LC50 1.24
Flow 20 150 NS 48-h LC50 4.9 Reiff et al.
(1979)
Flow 20 150 NS 48-h LC50 2.4
Flow 20 150 NS 48-h LC50 1.2
Flow 20 268 NS 48-h LC50 2.1-2.9
Flow 20 268 NS 96-h LC50 1.9-2.9
Flow 20 268 NS 48-h LC50 1.3-1.7
Flow 20 268 NS 96-h LC50 1.2-1.3
Flow 20 268 NS 48-h LC50 0.8-0.9
Flow 20 268 NS 96-h LC50 0.4-0.6
Himedaka (killifish) 4-5 wk Staticr 23 C8-C14 96-h LC50 10-18 Canton & Slooff
(Oryzias latipes) (1982)
323 mg Staticr 23-24 5.6-6.1 C11.7 48-h LC50 15 Kikuchi et
al.
323 mg Staticr 23-24 5.6-6.1 C11.7 48-h LC50 10 (1976)
approx. 262 mg Staticr 21-22 6.7-7.1 C12 48-h LC50 12 Kikuchi &
approx. 262 mg Staticr 21-22 6.7-7.1 NS 48-h LC50 10 Wakabayashi
(1984)
Table 28 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Himedaka (contd) C161 48-h LC50 > 50 Tomiyama (1974)
C6 48-h LC50 > 50
C8 48-h LC50 > 50
C10 48-h LC50 > 50
C12 48-h LC50 4
C14 48-h LC50 4 Iimori &
Takita
(1979)
25 7.2-7.9 48-h LC50 7.6 Hidaka et al.
(1984)
25 7.2-7.9 96-h LC50 7.3
Adult Staticr 20 5 C11.7 96-h LC50 13 Wakabayashi &
Adult Staticr 20 25 C11.7 96-h LC50 8.8 Onizuka (1986)
Adult Staticr 20 125 C11.7 96-h LC50 4.8
Adult Staticr 20 625 C11.7 96-h LC50 3.2
Adult Staticr 20 0 C11.7 48-h LC50 6.7 Wakabayashi &
Adult Staticr 20 1 C11.7 48-h LC50 4.8 Onizuka (1986)
Adult Staticr 20 5 C11.7 48-h LC50 4.7
Adult Staticr 20 10 C11.7 48-h LC50 3.5
Adult Staticr 20 15 C11.7 48-h LC50 3.8
Adult Staticr 20 20 C11.7 48-h LC50 2.5
Adult Staticr 20 25 C11.7 48-h LC50 1.9
Adult Staticr 20 30 C11.7 48-h LC50 1.6
Adult Staticr 20 35 C11.7 48-h LC50 1.4
Table 28 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Cobalt suzume 48-h LC50 1.3 Iimori &
(Chrysiptera hollisi) Takita (1979)
Smallmouth bass Flow 15 43 7.2-7.9 NS 96-h LC50 3.7 McKim et al.
(Micropterus dolomieu) (1975)
Black bullhead Flow 23 50 7.5 NS 96-h LC50 6.4 Thatcher &
(Ictalurus melas) Santner (1967)
Common shiner Flow 23 50 7.5 NS 96-h LC50 4.9 Thatcher &
(Notropis cornutus) Santner (1967)
Emerald shiner Flow 23 50 7.5 NS 96-h LC50 3.3 Thatcher &
(Notropis atherinoides) Santner (1967)
Bleeker 0.3 g Static 28 NS 96-h LC50 11.8 Eyanoer et al.
(Puntius gonionotus) 0.3 g Static 35 NS 96-h LC50 11.5 (1985)
0.3 g Static 50 NS 96-h LC50 13.6
0.3 g Static 110 NS 96-h LC50 11.8
0.3 g Static 260 NS 96-h LC50 11.4
Table 28 (contd)
Organism Size or Static or Temp. Hardness pH LAS chain End-point Concn Reference
age flow (°C) (mg/litre)a length (mg/litre)
Ayu 0.26 mg 1 NS 48-h LC50 0.86 Sueishi et al.
(Plecoglossus altivelis) 0.29 g 1 NS 48-h LC50 0.53 (1988)
1.24 g 1 NS 48-h LC50 0.77
6.51 g 1 NS 48-h LC50 1.45
27.98 g 1 NS 48-h LC50 1.17
Flow, flow-through conditions: LAS concentration in water maintained continuously; NS, not specified; staticr, static renewal: water changed
periodically; static, water unchanged for the duration of test; finger, fingerling
a mg/litre CaCO3
b Based on nominal concentration
c Based on measured concentrations
Table 29. Toxicity of linear alkylbenzene sulfonates (LAS) to marine fish
Organism Size or Static or Temp. Salinity LAS chain End-point Concn Reference
age flow (°C) (%) length (mg/litre)
Cod (Gadus morhua) 30 cm Static 6-8 32-34 C12 96-h LC50 1a Swedmark et al.
30 cm Static 15-17 32-34 C12 96-h LC50 < 1a (1971)
Flounder Static 6-8 32-34 C12 96-h LC50 1.5a
(Pleuronectes flesus) Static 15-17 32-34 C12 96-h LC50 < 1a
Plaice (Pleuronectes Static 6-8 32-34 C12 96-h LC50 > 1 -< 5a
platessa)
Mosbled sole Newly NS 48-h LC50 0.5-1 Yasunaga (1976)
(Limanda yokohamae) hatched
10 days NS 48-h LC50 0.1-0.5
30 days NS 48-h LC50 0.5-1
40 days NS 48-h LC50 < 0.1
Newly NS 48-h LC50 0.05-0.1
hatched
Olive flounder 5 days NS 48-h LC50 < 0.1
(Paralichtys olivaceus) 10 days NS 48-h LC50 0.1-0.5
30 days NS 48-h LC50 0.1-0.5
Himedaka (killifish) Adult Staticr 20 0 C11.7 48-h LC50 6.7 Wakabayashi &
(Oryzias latipes) Adult Staticr 20 1 C11.7 48-h LC50 4.8 Onizuka (1986)
Adult Staticr 20 5 C11.7 48-h LC50 4.7
Table 29 (contd)
Organism Size or Static or Temp. Salinity LAS chain End-point Concn Reference
age flow (°C) (%) length (mg/litre)
Himedaka (contd) Adult Staticr 20 10 C11.7 48-h LC50 3.5
Adult Staticr 20 15 C11.7 48-h LC50 3.8
Adult Staticr 20 20 C11.7 48-h LC50 2.5
Adult Staticr 20 25 C11.7 48-h LC50 1.9
Adult Staticr 20 30 C11.7 48-h LC50 1.6
Adult Staticr 20 35 C11.7 48-h LC50 1.2
Static: water unchanged for duration of test; staticr, static renewal: water changed periodically; NS, not specified
a Based on nominal concentration
The activity of the enzymes of aerobic metabolism was decreased, and
that of lactate dehydrogenase was strongly increased (Zaccone et
al., 1985). Concentrations of 2.19 mg/litre C11.8 LAS and
0.39 mg/litre C13 LAS significantly increased the 24-h mean
ventilation rate (number of opercular closures per minute) of
bluegill (Lepomis macrochirus) (Maki, 1979b).
A concentration of 36 mg/litre LAS severely affected the
viability of the perfused gills of rainbow trout (Oncorhynchus
mykissi). Vascular resistance increased gradually during
perfusion, with a concomitant decrease in oxygen transfer. LAS at
0.05 mg/litre more than doubled cadmium transfer (0.8 µg/litre)
through the perfused gills; and at concentrations of 36 mg/litre LAS
and 0.9 mg/litre cadmium, there was a marked reduction in cadmium
transfer (Pärt et al., 1985).
A9.3.3.4 Behavioural effects
Hidaka and co-workers have reported several studies of the
avoidance of surfactants by fish (Hidaka et al., 1984; Hidaka &
Tatsukawa, 1989; Tatsukawa & Hidaka, 1978). The results should be
interpreted with caution, since the environmental relevance and the
reproducibility and sensitivity of these tests is unclear;
furthermore, no effect was seen after removal of the olfactory
organs. Another study (Maki, 1979a) showed no adverse toxicological
effects at concentrations two times greater than those reported to
cause avoidance reactions.
Hidaka et al. (1984) found that the minimal avoidance
concentration of LAS, i.e. the concentration at which fish spent 65%
of a 5-min period in clean water in order to avoid LAS, was
13.5 µg/litre for medakas (Oryzias latipes). Medakas exposed to
LAS at concentrations of 5-50 µg/litre for 10 min showed significant
avoidance to 10, 20, and 30 µg/litre. No significant avoidance of
concentrations of 10-50 µg/litre LAS was found after removal of the
olfactory organs (Hidaka & Tatsukawa, 1989).
The threshold concentrations for avoidance of LAS by ayu
(Plecoglossus altivelis) were 0.11 µg/litre of a formulation and
1.5 µg/litre of pure reagant LAS (Tatsukawa & Hidaka, 1978).
A9.3.3.5 Interactive effects with other chemicals
The chronic toxicity of para, para-DDT (50 mg/litre) to
goldfish (Carassius auratus) was increased by prior exposure to
LAS at 4 mg/litre for 37 days (Dugan, 1967).
The toxicity of 1 mg/litre LAS solution to mosquito fish exposed
under static conditions was not affected by allowing the LAS
solution to react with excess chlorine (Katz & Cohen, 1976).
A concentration of 1 mg/litre LAS significantly increased the
toxicity of fuel oil to bluegill (Lepomis macrochirus), reducing
the 24-h LC50 from 91 to 51 mg/litre. The authors concluded that
sublethal concentrations of LAS increased the toxicity of fuel oil
by increasing its penetration (Hokanson & Smith, 1971). The toxicity
of No. 2 and No. 4 fuel oils in six species of freshwater fish was
increased in the presence of 1-5 mg/litre LAS (Rehwoldt et al.,
1974).
The uptake of cadmium by freshwater trout (Salmo gairdneri)
exposed to 0.14 µmol/litre LAS was more than two times greater than
in controls. Reduced cadmium uptake was reported in fish exposed to
100 µmol/litre LAS. The authors reported that trout exposed to low
levels of both LAS and cadmium could take up lethal cadmium
concentrations. LAS were reported to interact with the gill proteins
involved in cadmium transport, resulting in increased permeability
to cadmium (Pärt et al., 1985).
Fathead minnows (Pimepheles promelas) were exposed to various
pesticides in the presence and absence of 1 mg/litre LAS. Parathion
acted synergistically with LAS, reducing the 96-h LC50 from 1410
to 720 µg/litre. Endrin and LAS showed no synergism, and no
consistent results were obtained for DDT (Solon et al., 1969).
Methyl parathion, ronnel, trithion, and trichloronat were also found
to act synergistically with LAS, but neither ortho-ethyl-ortho-4-
nitrophenyl phenylphos-phonothioate nor dicapthon exhibited
synergism. The synergistic relationship does not appear to be
exclusive to a general structural group (Solon & Nair, 1970).
Goldfish (Carassius auratus) were exposed to mixtures of LAS
and chloramines and LAS and copper at ratios of 1:1, 2:1, and 1:2,
and toxicity curves and 24-h and 96-h LC50 values were compared.
LAS and chloramines had an additive effect at a ratio of 1:1, but at
2:1 and 1:2 synergistic effects were seen. LAS and copper at ratios
of 1:1 and 2:1 had additive effects; however, at 1:2, high
concentrations and longer exposure times had additive effects, and
low concentrations and shorter exposure times had synergistic
effects (Tsai & McKee, 1978).
When eggs of cod (Gadus morhua) were exposed to mixtures of
LAS and zinc or copper from fertilization to hatching, zinc had a
weak synergistic affect on the toxicity of LAS, but LAS had a strong
synergistic affect on the toxicity of copper (Swedmark & Granmo,
1981).
In a study of the effect of polyoxyethylene (20) on the acute
toxicity of C12 LAS, red killifish (Oryzias latipes) and carp
(Cyprinus carpio) were exposed to the 48-h LC50 of LAS for the
respective species and to 5-40 mg/litre of either a polyoxyethylene
sorbitan ester, a polyethylene glycol, a polypropylene glycol, or a
protein (albumin, kaolin, or bentonite). Addition of most of these
substances decreased mortality. No mortality was observed in carp
exposed to LAS and 14 or 28 mg/litre polyoxyethylene (20) sorbitan
monooleate (SMOE20) or to other nonionic surfactants with a similar
polyoxyethylene sorbitan ester structure-polyoxyethylene (6)
sorbitan monolaurate, polyoxy-ethylene (6) sorbitan monooleate,
polyoxyethylene (20) sorbitan monolaurate, and polyoxyethylene (20)
sorbitan monostearate-or to albumin. No significant effect on
mortality induced by LAS was reported after simultaneous exposure to
polyoxyethylene (6) sorbitan monostearate, polyethylene glycol,
polypropylene glycol, kaolin, or bentonite. The authors also
examined the histological effects of these chemicals on the gills of
carp exposed to high concentrations of LAS, including the 48-h
LC50 of 3.5 mg/litre and the LC100 of 7 mg/litre. Histological
changes in fish exposed only to 3.5 mg/litre LAS included the
appearance of mucous cells and agglutination of the secondary
lamellae; carp exposed to a mixture of LAS and SMOE20 showed only
slight swelling of the secondary lamellae and slight proliferation
of the gill epidermal cells. Exposure only to LAS at 7 mg/litre
resulted in marked proliferation of the epidermal cells and
agglutination of secondary lamellae; exposure to both LAS and SMOE20
induced only swelling of the secondary lamellae. No effects were
reported on the gills of control fish or on other organs of the
exposed fish; and no significant differences from controls were
reported in haematological or serum biochemical parameters for fish
exposed to either LAS or the LAS:SMOE20 mixture. When the
distribution of LAS in the tissues and organs of carp was examined,
higher levels were found in the blood and most organs after exposure
to LAS only than after exposure to the mixture; the differences were
statistically significant in blood, muscle, and gill but not in
spleen or gall-bladder. Adsorption of the 5- and 6-phenyl isomers of
LAS was similar when they were given alone or in conjunction with
SMOE20, but more of the 4- and (especially) the 2-phenyl isomers was
adsorbed by fish receiving LAS alone, indicating that SMOE20
decreases the acute toxicity of LAS to fish by decreasing the
adsorption on the gills of the more toxic isomer (Toshima et al.,
1992).
A9.3.4 Amphibia
No reliable data were available.
A9.3.5 Studies of the mesocosm and communities
Diversity and similarity indices were used in many studies to
assess the effects of LAS on phytoplankton communities, usually on
the basis of taxonomy, mean number of species, and density. Mean
density and similarity indices were then compared with those of
controls. In general, these indices are not sensitive to change, as
the densities of some species may decrease while the indices do not.
The effects of C12 and C13 LAS on short-term photosynthetic
activity were studied in plankton sampled from Acton Lake, Ohio,
United States, during May-October in the laboratory and in situ.
Toxicity increased with the temperature of the water, the most
sensitive period being June-August, and LAS were less toxic during
periods of diatom dominance and low phytoplankton density. Thus the
density of diatoms decreased during June-August, and that of green
and blue algae increased. The comparison of the results of
laboratory and field tests was highly dependent on species and the
in-situ end-point. Short-term tests for photosynthetic activity in
situ gave 3-h EC50 values of 0.2-8.1 mg/litre (mean, 1.9 mg/litre)
for C13 LAS and 0.5-8.0 mg/litre (mean, 3.4 mg/litre) for C12
LAS (Lewis & Hamm, 1986). (See also section 9.3.1.1.)
In another study of the effect of LAS on phytoplankton
communities in Acton Lake, Ohio (Lewis, 1986), phytoplankton were
exposed in situ to LAS at a concentration of 0.01, 0.02, 0.24,
0.80, 27, or 108 mg/litre for 10 days. The LOEL for LAS, based on
community similarity indices and the mean number of species, was
108 mg/litre. The similarity index (coefficient of community)
decreased as the concentration of LAS increased, with calculated
values of 0.62 at 0.01 mg/litre and 0.43 at 108 mg/litre. No
significant effects were seen on the community diversity index or
phytoplankton density. Green algae were the species least affected,
on the basis of abundance, followed in order of decreasing tolerance
by blue-green algae and diatom species. Chlorophyta species were
the most abundant at higher concentrations of LAS, comprising 74% of
the total cell volume at 108 mg/litre; their abundance tended to
increase to a maximum at this concentration and then decrease to
values similar to those of the controls. Chlorophyta species of
the genera Chlamydomonas, Oocystis, and Sphaerocystis were not
found after exposure to higher concentrations of LAS. Chlamydomonas
was found only in waters with a concentration of LAS <
0.8 mg/litre, and Oocystis and Sphaerocystis were found only at
concentrations < 27 mg/litre. The peak density of blue-green
phytoplankton (56% of cell volume) was achieved at 0.24 mg/litre
LAS, declining to 17% at 108 mg/litre. The density of the major
species, Schizothrix calcicola, was greatest at 27 mg/litre LAS but
declined to a level below that of controls at 108 mg/litre LAS. The
abundance of diatoms was low at all concentrations of LAS. At
concentrations < 0.24 mg/litre, the average density of diatoms
was 23% of the total cell volume, similar to that of controls; at
concentrations of 0.24-0.8 mg/litre, the diatom density was 10% of
the cell volume. The mean densities of the major diatom species,
such as Cyclotella glomerata, Cyclotella pseudostelligera, and
Nitzschia frustulum v. perminuta, followed the overall trend for
diatoms, reaching a peak at low LAS concentrations and declining to
control values at higher concentrations.
In the same study, the laboratory-based 96-h EC50 values for
exposure to C11.8 LAS were calculated to be 29.0 mg/litre for
Selenastrum and 0.0096 mg/litre for Microcystis, on the basis of
population growth. The lowest concentration of LAS that produced a
significant effect on algal growth in the laboratory was
0.05-1.0 mg/litre, which is considerably lower than the
27-108 mg/litre value found to be the lowest that altered the
structure of a natural phytoplankton community. The differences
between the results of laboratory and field tests were smaller for a
laboratory-based EC50 than for an LOEL. Calculations based on the
EC50 give a 30-fold difference for Microcystis but essentially no
difference for Selenastrum (Lewis, 1986).
The toxic effects of LAS were also examined on periphyton
communities above and below the outfall of a wastewater treatment
plant on Little Miami River, Ohio, United States. The dominant
species at both test sites were diatoms, Amphora perpusilla and
Navicula minima accounting for at least 80% of the total cell
volume. The tests were conducted in situ, with 21-day
continous-flow exposure to LAS (average chain length, C11.9) in
river water entering submerged exposure tubes at a concentration of
0.2, 1.1, 9.8, or 28.1 mg/litre, after a four-week colonization
period. The delivery rate of LAS was adjusted daily according to
measurements of river flow in order to maintain the desired test
concentrations. The periphyton at the site below the treatment plant
outfall were exposed to LAS in the presence of 20-30% treated
municipal effluent. No effects on the structure or function of the
periphyton community above the outfall were reported after exposure
to an average concentration of LAS < 1 mg/litre. The lowest
concentration that had an effect on the upstream periphyton
community was 9.8 mg/litre, which reduced photosynthesis by 16%; a
concentration of 28.1 mg/litre reduced photosynthesis by 64%, with a
noticeable reduction in chlorophyll a. No effects on community
similarity or diversity were reported in comparison with control
communities. Mean cell densities were increased by 26% after
exposure to 0.2 mg/litre LAS and by 17% after exposure to
1.1 mg/litre; exposure to 28.1 mg/litre reduced mean cell density by
28%. Exposure to LAS had no significant effects on the abundance of
the three main species in the upstream community. Increased
photosynthesis (by 12-39%) and chlorophyll a (50-51%), were
reported after exposure to 1.1 or 9.8 mg/litre LAS, but exposure to
28.1 mg/litre resulted in a 52% decrease in photosynthesis and a 71%
decrease in chlorophyll a. No effects on the similarity or
diversity of the periphyton community were reported at any
concentration of LAS tested. Cell densities of periphyton were
increased by 34% after exposure to 9.8 mg/litre LAS and by 13% after
exposure to 28.1 mg/litre. The abundance of the three main species
in the downstream periphyton community was not affected. The lowest
concentration of LAS that induced an effect was 3.3 mg/litre for the
upstream periphyton community and 16.6 mg/litre for the downstream
community. The authors suggested that the difference between the two
values was due to the presence of 20-30% sewage downstream, which
reduced the bioavailability of LAS (Lewis et al., in press).
When an aquatic ecosystem was exposed to LAS at concentrations
of 0.25-1.1 mg/litre for 90 days, the numbers of phytoplankton were
unaffected but primary productivity was significantly reduced at all
concentrations. The zooplankton population showed a more variable
response: the number of rotifers was reduced at all concentrations,
and those of Diaptomus and Cyclops were reduced at >
0.51 mg/litre. The number of ostracods was decreased at
0.38 mg/litre but was increased at 0.51 and 1.1 mg/litre. The
chironomid population was significantly reduced at concentrations
> 0.38 mg/litre (Chattopadhyay & Konar, 1985). Exposure of an
aquatic ecosystem consisting of phytoplankton, zooplankton, and
benthic organisms to 1 mg/litre of a preparation of LAS for 90 days
significantly reduced the numbers of phytoplankton and zooplankton
per litre but did not significantly affect the numbers of chironomid
larvae (Panigrahi & Konar (1986).
The effect of C12 LAS on microbial communities was studied in
a model ecosystem consisting of a 19-litre glass tank containing
sediment from Winton Lake, Ohio, United States, and several trophic
levels, comprising bacteria, algae, macrophytes (Elodea canadensis,
Lemna minor), macroinvertebrates (Daphnia magna, Parantanytarsus
parthenogenica), and blugill sunfish (Lepomis macrochirus).
After a four-week equilibrium period, LAS were added at 0.5 or
5.0 mg/litre to a flow-through system with six to 10 replacements
per day for 26 days. The structure of the microbial communities was
not affected, and no differences were reported in mean biomass or
number of colony-forming units between the microorganisms exposed at
the two levels. The function of the microbial communities, assayed
by measuring the degradation of both LAS and D-glucose, was reduced
only at 5.0 mg/litre. In a similar system, in which the same
concentrations of LAS were added in the form of sewage effluent, no
effect was seen on the structure of the microbial community or on
their function, measured only as the degradation of LAS (Larson &
Maki, 1982).
Addition of LAS (average chain length, C11.8) at a measured,
relatively uniform concentration of 0.36 ± 0.05 mg/litre to 50-m
outdoor experimental streams had no effect on total density, species
richness, percentage similarity, or dominance of macroinvertebrates
or periphyton or on the processing of organic matter of leaf discs.
Fathead minnows (Pimephales promelas) and amphipods (Hyallela
azteca) were exposed in groups of 10 and 20 per box placed in the
streams at three locations. The mortality rates of the amphipods
were 17-25% after exposure to LAS and 47% among controls; no effects
were seen on the survival or weights of the fish, although minor
effects were found on length (Fairchild et al., 1993).
A study of the fate and effects of surfactants in outdoor
artificial streams addressed the effect of LAS on drift and
population densities of macroinverebrates, the reproductive
behaviour of an amphipod, the scud (Gammarus pulex), the survival
of a fish, the three-spined stickleback (Gasterosteus aculeatus),
and photosynthesis by the community. The concentration of LAS in
sediment was reported to increase with increasing water
concentration, and selective adsorption of longer-chain LAS
homologues to sediment was reported. The microbial populations of
both the water and the sediment adapted to LAS, resulting in a
reduction in its half-life during the test. LAS at concentrations
< 1.5 mg/litre did not affect macroinverebrate drift, population
density, or community photosynthesis. Survival of the fish and
reproduction by the amphipod were affected at concentrations of
1.5-3.0 mg/litre (Mitchell & Holt, 1993).
A9.3.6 Field studies
The effect of LAS downstream of a sewage outflow was studied by
monitoring sediment, water, and the distribution of invertebrates at
an upstream control site, a site near the discharge point, and a
site 200 m downstream of the outflow. The concentrations of LAS in
sediment were 1-40 mg/kg dry weight, with concentrations < 2 mg/kg
at the control site and 200 m downstream. No effect of LAS in the
effluent or in the streambed sediments could be discerned on the
invertebrate populations (Ladle et al., 1989).
A9.3.7 Toxicity of biodegradation intermediates and impurities
of linear alkylbenzene sulfonates
Tests of degradation products and impurities of LAS show that
they are less toxic than LAS themselves.
A9.3.7.1 Individual compounds
The 48-h LC50 values in Daphnia magna were 208 ± 85 mg/litre
for sulfophenylundecanoic acid, disodium salt (mixed isomers, 6-10
phenyl); about 6000 mg/litre for 3-(sulfophenyl) butyric acid,
disodium salt; and about 5000 mg/litre for 4-(sulfophenyl) valeric
acid, disodium salt. The equivalent 48-h LC50 values in the
fathead minnow (Pimephales promelas) were 77 ± 12, about 10 000,
and about 10 000 mg/litre, respectively (Kimerle & Swisher, 1977).
The 24-h LC50 values in Daphnia were about 22 000 mg/litre for
3-sulfophenylbutyric acid, disodium salt; about 12 000 mg/litre for
3-sulfophenylheptanoic acid, disodium salt; > 22 000 mg/litre for
3-sulfophenylbutyric acid, disodium salt; and 2 000 mg/litre for
sulfophenylundecanoic acid, disodium salt. Other tests were carried
out with the last two compounds, giving 96-h LC50 values of about
28 000 and 1200 mg/litre in fathead minnows (Pimephales promelas);
28-day NOELs of > 2000 and > 200 mg/litre for survival and
reproduction of Daphnia; and 30-day NOECs of > 1400 and >
52 mg/litre for survival of the fry of fathead minnows (egg
hatchability and fry growth were less sensitive) (Swisher et al.,
1978).
The 96-h LC50 for mixed isomers of sulfophenylundecanoic acid
disodium salt in bluegill (Leponis macrochirus) was 75 mg/litre
(Swisher et al., 1964). The 24-96-h LC50 values in fathead minnows
were 1000-1500 mg/litre for sulfophenylundecanoic acid (C11) and
25 000-32 000 mg/litre for sulfophenyl butyrate (C4) (Swisher et
al., 1978).
The 48-h LC50 for the alkanoic acid derivatives of
2-sulfophenyl C13 LAS and 4-sulfophenyl C13 LAS in nearly pure
form was > 800 mg/litre in goldfish (Carassius auratus) (Divo &
Cardini, 1980).
The 24-h LC50 values for Daphnia magna exposed to
dialkyltetralin sulfonates, which are trace contaminants of LAS,
were 420, 195, 110, 50, and 27 mg/litre for tetralin sodium
sulfonates of chain lengths C10, C11, C12, and C13,
respectively (Arthur D. Little Inc., 1991).
A9.3.7.2 Effluents
Interpretation of tests on effluents must take into account the
following:
-- As concentrations arwe often reported as MBAS, testing of
effluent from a sewage treatment plant may result in overestimation
of the actual concentrations of LAS, owing to interference (see
section 2.3).
-- The bioavailability of LAS is decreased by the presence of high
concentrations of suspended solids; thus, as effluents are diluted
in the environment, availability is usually increased, although
biodegradation occurs.
Addition of LAS (C10-C15) to detergent-free activated sludge
plant effluent (95% was removed as MBAS) gave a nominal 96-h LC50
in rainbow trout (Oncorhynchus mykiss) of 0.36 mg/litre. After
treatment, the 96-h LC50 was 29.5 mg/litre, expressed in terms of
the concentration of the surfactant in the influent (Brown et al.,
1978).
When bluegill were exposed to effluent from continuous-flow
activated sludge units fed 100 mg/litre LAS, none died during
4-11-day exposure (Swisher et al., 1964).
A9.4 Terrestrial organisms
A9.4.1 Terrestrial plants
Young seedlings of tomato, lettuce, radish, pea, cucumber, and
barley were grown in a soil-based compost and were watered and given
a foliar spray of a preparation of LAS. No effects were noted at
concentrations up to 100 mg/litre (Gilbert & Pettigrew, 1984). In
another study, barley, tomato, and bean plants were grown from seed
and watered with a solution containing LAS at a concentration of 10,
25, or 40 mg/litre. Plants that received the lowest dose germinated
at the same time as controls, but plants watered at 25 or
40 mg/litre germinated three days later. The growth of barley plants
was inhibited at all three concentrations; however, the dose of
25 mg/litre increased the growth rate of beans, and the highest dose
increased the growth rate of both tomatoes and beans (Lopez-Zavala
et al., 1975).
The 21-day EC50 values for LAS (C10-C13), based on the
emergence of seedlings and early stages of growth, were 167 mg/litre
in sorghum, 289 mg/litre in sunflower, and 316 mg/kg in mung bean.
The highest concentration that caused no significant reduction in
the growth of any of the three species was 100 mg/kg (Holt et al.,
1989; Mieure et al., 1990). In a second study, 407 mg/kg C11.36 or
393 mg/kg C13.13 LAS were mixed with sewage sludge, and nine
common plant species, including five crop plants, were exposed as
seed either at the same time or two weeks after application of the
sludge to soil at a rate of 9000 kg/ha. There was no significant
effect on seed germination and no significant inhibition of growth
(Mieure et al., 1990).
Orchid seedlings (Phalaenopsis or Epidendrum sp.) were grown
in culture media containing either the sodium or the ammonium salt
of LAS at a concentration of 10, 100, or 1000 mg/litre. The lowest
dose had no effect on growth, and that of ammonium LAS had no effect
on germination. At 100 mg/litre, survival was halved and germination
completely inhibited (Ernst et al., 1971). A concentration of
1000 mg/litre caused drastic changes in morphology, loss of
membranes, swelling of thylakoids, and the appearance of dense
osmophilic granules in chloroplasts (Healey et al., 1971).
The growth of pea seedlings grown for 26 days in quartz sand to
which 0.005% (50 mg/kg) LAS had been added was significantly
reduced, as measured by the fresh weight of roots and the length and
fresh weight of pea greens (Lichtenstein et al., 1967).
LAS were not toxic with respect to growth at the early life
stages of radish, Chinese cabbage, and rice when added in hydroponic
culture at concentrations of 10, 20, and 20 mg/litre, respectively;
concentrations of 20, 35, and 35 mg/litre were toxic (Takita, 1982).
When seeds of Pisum sativum and Crotolaria juncea were
exposed to LAS for 24 h before sowing, the percentage germination
was reduced at concentrations of 1 ml/litre for P. sativum and
10 ml/litre for C. juncea, although no statistical analysis was
presented. No germination occurred after exposure to LAS at
concentrations of 20 ml/litre for P. sativum and 40 ml/litre for
C. juncea. Radicle length was reduced at > 0.1 ml/litre in both
species (Sharma et al., 1985).
Application of LAS at 50 g/m2 under field conditions to loamy
and sandy soils (corresponding to 0.47-1 mg/kg dry weight,
respectively) led to considerable physiological damage, including
leaf necrosis, chlorosis, and turgescence, to ryegrass (Lolium
perenne) after 14 days; however, there was no difference in the
fresh weight yield after harvesting at 45-54 days (Litz et al.,
1987).
A9.4.2 Terrestrial invertebrates
When the earthworm Eisenia foetida was exposed to C11.36 LAS
incorporated into soil at nominal concentrations of 63-1000 mg/kg
dry weight, the 14-day LC50 was > 1000 mg/kg. On the basis of a
statistical analysis of body weights, the no-effect concentration
was 250 mg/kg; this was confirmed by HPLC to be 235 mg/kg. In a
second study, C11.36 and C13.13 LAS were incorporated into
sludge and applied to soil, and the earthworm Lumbricus terrestris
was exposed to the subsequent mixture, which contained LAS at
concentrations of 84-1333 mg/kg. The 14-day LC50 was again found
to be greater than the highest concentration (> 1333 mg/kg). The
no-effect concentration, based on weight and burrowing behaviour,
was the nominal concentration of 667 mg/kg, measured by HPLC as
613 mg/kg. The worms were exposed, however, to LAS under conditions
of continuous light, which would inhibit them from surfacing to feed
and thus increase their exposure to and the toxicity of the test
over that of the same concentration in the field (Mieure et al.,
1990).
Topical application to house flies (Musca domestica) of LAS at
the same time as parathion, diazinon, or dieldrin in ratios of 1:1
and 1:10 had no effect on the toxicity of the insecticides. When LAS
were added to soil treated with parathion or diazinon, however, a
significant synergistic effect was observed on the toxicity of the
insecticides to the fruit fly Drosophila melanogaster. The optimal
concentration of LAS that resulted in synergy was 23 mg/kg
(Lichtenstein, 1966).
A9.4.3 Birds
No significant effect on egg quality was found after Leghorn
chickens were fed a diet containing 200 mg/kg LAS for 45 days
(Lopez-Zavala et al., 1975).
B. alpha-Olefin sulfonates
B1. SUMMARY
See Overall Summary, Evaluation, and Recommendations (pp. 7-21).
B2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND
ANALYTICAL METHODS
B2.1 Identity
Chemical formula: CnH2nO3S Na, CnH2n+1O4S Na ( n = 14-18)
Chemical structure: CH3(CH2)jCH:CH(CH2)kSO3- Na+
CH3(CH2)mCH(CH2)nSO3- Na+
OH ( m,n, integers)
Common names: Sodium alpha-olefinsulfonate,
alpha-olefin-sulfonic acid sodium salt, AOS
sodium salt
Common trade names: Bioterge AS 40 F, Elfan OS 46, Geropon
MLS/A, Hostapur OS Brands, Lipolan, Lipomix
G, Lipon PB-800, Lutensit A-PS, Nansa
LSS38/AS, Sawaclean, Sermul EA 214,
Sulframin AOS, Witconate (McCutcheon, 1989)
Abbreviations: AOS, AOS-Na
CAS Registry numbers: 29963-33-5 Sodium 1-tetradecenesulfonate
29734-60-9: Sodium hexadecenesulfonate
13513-23-0: Sodium 3-hydroxyhexadecyl-1-
sulfonate
26446-92-4: Octadecene-1-sulfonic acid
sodium salt
13513-42-3: 3-Hydroxy-1-octadecanesulfonic
acid, sodium salt
Specifications: AOS are mixtures consisting of about 60-65%
alkene sulfonates, 30-35% hydroxylalkane
sulfonates, and 5-10% disulfonates. Various
positional isomers of alkene sulfonates and
hydroxyalkane sulfonates have been reported
(Gentempo et al., 1985; Williamson, 1993).
Sodium C14-C16 AOS are typically
shipped as solutions containing 35-40%
active matter in water. Sodium C16-C18
AOS are typically slurries containing
28-30% active matter in water at ambient
temperature.
B2.2 Physical and chemical properties
AOS are white crystalline solids consisting of various chemical
compounds and their isomers, with different properties. Typical
properties of AOS are given in Table 30. Two ranges are usually
offered; the commonest are based on C14-C16 olefin and the other
on C16-C18 olefin. Detergency is maximal with alkyl chain
lengths of C15-C18. Maximal detergency is also obtained with the
same range of alkyl chain lengths in a detergent formulation that
includes alkali builders and chelating agents (Yamane et al., 1970).
AOS are stable, even in hot acidic media.
Table 30. Relationship between alkyl chain length, Krafft point,
critical micelle concentration (CMC), and surface
tension of alpha-olefin sulfonates
Alkyl chain Krafft pointa CMCal Surface tension
length (°C) (g/litre) (dyne/cm)
12 - 4.0 -
14 - 1.0 30
16 10 0.3 33
18 30 0.1 35
20 40 - -
(25°C) (25°C)
From Ohki & Tokiwa (1970)
a The solubility of surfactants in water, defined as the
concentration of dissolved molecules in equilibrium with a
crystalline surfactant phase, increases with rising temperature. For
surfactants, there is a distinct, sharp bend (break-point) in the
solubility-temperature curve. The steep increase in solubility
above the sharp bend is caused by micelle formation. The point of
intersection of the solubility and critical micelle curves, plotted
as a function of temperature, is referred to as the Krafft point.
This is a triple point at which surfactant molecules coexist as
monomers, micelles, and hydrated solids. The temperature
corresponding to the Krafft point is called the Krafft temperature.
At above the Krafft temperature and critical micelle concentration,
a micellar solution is formed. Under these conditions, higher levels
than the aqueous solubility may be obtained.
B2.3 Analytical methods
There is no officially recognized specific procedure for the
analysis of AOS in environmental samples. The methods commonly used
to analyse anionic surfactants are also used for AOS, except those
involving high-performance liquid chromatography (HPLC), which has
limited use in environmental analyses for AOS, because they do not
absorb ultra-violet radiation as effectively as do linear
alkylbenzene sulfonates (LAS). A modified version of the methylene
blue-active substance (MBAS)-HPLC method described in the monograph
on LAS has been developed (Takita & Oba, 1985).
Nonspecific methods used in the analysis of anionic surfactants
in general, such as the MBAS method, can be used to analyse
materials for AOS (see section 2.3 of the monograph on LAS).
B3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE
B3.1 Natural occurrence
AOS do not occur naturally.
B3.2 Anthropogenic sources
B3.2.1 Production levels and processes
AOS are synthesized industrially. Although they have been
available since the 1930s, production for use in commercial
surfactant formulations was somewhat limited until recently owing to
a lack of suitable feedstock. Development of continuous and
short-contact sulfur trioxide sulfonation processes and the
increased availability of highly pure Ziegler-derived alpha-olefin
feedstock has recently made AOS surfactants competitive with other
surfactants on the market (Arthur D. Little Inc., 1977, 1981).
The estimated world consumption of AOS in 1988 was 50 200 tonnes
(Colin A. Houston & Associates Inc., 1990). In 1990, that group
estimated that world consumption would be 51 900 tonnes; an
alternative estimate (Hewin International Inc. 1992) was 76 000
tonnes (Table 31).
Table 31. Estimated worldwide consumption of alpha-olefin
sulfonates (tonnes)
Region Household Personal Industrial and All uses
products care institutional
products use
North America 3 000 7 000 4 000 14 000
Western Europe 2 000 3 000 3 000 8 000
Japan 24 000 7 000 2 000 33 000
Rest of the 18 000 3 000 - 21 000
world
Total 57 000 20 000 9 000 76 000
From Hewin International Inc. (1992)
AOS are prepared commercially by direct sulfonation of linear
alpha-olefins with a dilute stream of vaporized sulfur trioxide in a
continuous thin-film reactor. The olefin is obtained by wax cracking
or ethylene polymerization with a Ziegler-type catalyst (Tomiyama,
1970). The reaction is complex and follows several paths, forming
large amounts of various sultones as intermediates which hydrolyse
during subsequent quenching and neutralization. Commercial AOS
products contain a mixture of two major components, alkene sulfonate
and hydroxyalkane sulfonate, with smaller amounts of alkene
disulfonates, hydroxyalkane disulfonates, and saturated sultones.
B3.2.2 Uses
AOS are good detergents, have good foaming characteristics in
hard water and are used in heavy-duty laundry detergents, light-duty
dishwashing detergents, shampoos, and cosmetics. Table 31 indicates
the use patterns for AOS.
B4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
Section summary
It can be inferred that AOS are transported into the environment
in a similar manner to that established for LAS, alkyl sulfates, and
other detergent surfactants. Fewer data are available on the
environmental transport, distribution, and transformation of AOS
than for LAS. The environmental fate of AOS is similar to that of
LAS and alkyl sulfates: it is readily biodegraded under aerobic
conditions, and primary biodegradation is complete within 2-10 days,
depending on the temperature. At temperatures below 5-10°C,
biodegradation kinetics are reduced, owing to a reduction in
microbial activity. No data were available on abiotic degradation.
There was no evidence of bioaccumulation or bioconcentration in a
study of fish in which the uptake and distribution of AOS were
examined.
B4.1 Transport and distribution between media
In the same manner as other detergent compounds, AOS are
discharged into the environment in wastewater. The wastewater may
undergo sewage treatment if such facilities are available. In
countries where there are no adequate wastewater treatment
facilities, AOS released to the environment are removed by
biodegradation and adsorption mechanisms (see section 4.2 of the
monograph on LAS).
Limited studies of the adsorption of AOS are available. In a
study of the adsorption of C12 AOS on river sediments, the
equilibrium quantities adsorbed were proportional to the organic
carbon content of the sediments, with a sorption coefficient Koc
(dimensionless; normalized for the level of organic matter) of 0.65.
This indicates that adsorption of C12 AOS is slightly weaker,
than, for example, that of C12 LAS or C12 alkyl sulfonates
(Urano et al., 1984). Like other detergent chemicals, AOS are
adsorbed onto sewage sludge and river sediments in the environment.
B4.2 Biotransformation
B4.2.1 Biodegradation
B4.2.1.1 Aerobic biodegradation
Primary biodegradation of AOS, studied in die-away tests in
water from various sites on the Tama River, Japan, was complete
within three to five days when measured by the MBAS method; however,
total organic carbon was completely removed after an incubation time
of 20 days. In a study of AOS in seawater collected from the mouth
of the Tama River, 99% of MBAS was removed within one day, and 90%
of organic carbon was removed within five days (Sekiguchi et al.,
1975b).
In a comparison of the MBAS and total organic carbon methods for
measuring biodegradation with the shake-culture method, AOS lost 99%
of their activity as measured by the MBAS method and 90% of total
carbon within one day; 100% was lost within five days (Sekiguchi et
al., 1975a). In another study, complete loss of parent AOS (initial
concentration, 100 mg/litre) as determined by the MBAS method was
seen within 15 days, and 90% of total organic carbon was removed
within eight days (Miura et al., 1979). In a static die-away test
system, 90% biodegradation of three commercial AOS products,
comprising 100% C14-C16 AOS and > 95% C15-C18 AOS (determined
as MBAS), was reported within four days (Gafa & Lattanzi, 1974).
In a shake-culture test in Bunch-Cambers medium, C15-C18 AOS
were degraded by 99% (determined as MBAS) or 90% (removal of total
organic carbon) within one day; 100% total organic carbon was
removed within five days. The authors did not verify whether the
removal was the result of adsorption or mineralization (Sekiguchi et
al., 1972). The biodegradation of C15 AOS and three C15-C18
compounds with disulfonate contents of < 4, 15, and 50% in a
shake-flask culture system was reported to be 96% (determined as
MBAS), with no significant difference between compounds (Oba et al.,
1968b).
In a modified OECD screening test, 85% of C14-C18 AOS
(measured as chemical oxygen demand) was removed. Measurement of
MBAS in the same test indicated 99% removal (Gerike, 1987).
The aerobic biodegradation of 20 mg/litre AOS at 27°C was
followed during a 10-day incubation period. Primary degradation,
measured by the MBAS method, was complete within 10 days. The
theoretical CO2 production had reached 30-40% within that time
(Itoh et al., 1979).
The oxygen uptake of C14-C18 AOS was reported to be 85% of
the theoretical oxygen demand in a closed-bottle test (Gerike,
1987). The average biochemical oxygen demand for C12-C18 AOS
containing up to 40% hydroxylalkane sulfonates was 51.6% at five
days, while glucose under the same conditions had a biochemical
oxygen demand of 69.6% (Procter & Gamble Co, unpublished data).
The primary and ultimate biodegradability of a series of pure
AOS homologues (C12, C14, C16, and C18) was determined by
measuing CO2 production. Primary biodegradation was 98-99% within
three days, the rate of degradation varying with chain length.
Degradation of C12 and C14 AOS occurred at a similar rate (65%
within 30 days), but C18 AOS degraded more slowly. Mineralization
of all AOS samples was reported to be at least 50% within two weeks,
whereas mineralization of glucose during that time was 75-80%
(Kravetz et al., 1982). In a study of the biodegradation of the two
major breakdown products of AOS, alkene sulfonate and hydroxyalkane
sulfonate, AOS homologues (C15, C16, C17, C18) were degraded
to about 50%, and in each case the alkene sulfonate was degraded at
least twice as fast as the hydroxyalkane sulfonate (Sekiguchi et
al., 1975c).
The biodegradation of C18 AOS at a concentration of
28 mg/litre was studied in activated sludge (concentration, 100 mg
dry matter per litre) over 12 days: 90% was lost within eight days,
as measured by removal of chemical oxygen demand. The specific rate
of biodegradation was calculated to be 5.3 mg/g per h (Pitter &
Fuka, 1979).
In the OECD confirmatory test with activated sludge, 20 mg/litre
AOS were degraded, as follows: 97% C14 AOS within 17 days, 98%
C16 AOS within seven days, and 94% C14-C18 AOS within eight
days (Maag et al., 1975).
Primary biodegradation of C15-C18 AOS was dependent on
incubation temperature in die-away tests with water from the Tama
River, Japan. Primary biodegradation was complete within two days at
27°C, within five days at 15°C, and within two days at 21°C;
however, at a water temperature of 10°C about 20% of the AOS had
still not been degraded within the nine-day test (Kikuchi, 1985).
When C15-C18 AOS were added to seawater, no MBAS activity
was present after five days (Marquis et al., 1966).
B4.2.1.2 Anaerobic degradation
The primary anaerobic biodegradation of C15-C18 AOS
(measured as MBAS) by bacteria on sludge sampled from a sewage
treatment plant was 19% within 14 days and 31% within 28 days. More
parent AOS were degraded by bacteria from the bottom of a private
cesspool, with 34% lost within 14 days and 43% within 28 days. The
anaerobic degradation reported may have been due to the presence of
hydroxyalkane sulfonate compounds (Oba et al., 1967). AOS and LAS
were reported to be the two surfactants that were least degraded
anaerobically (Itoh et al., 1987).
B4.2.2 Abiotic degradation
No information was available.
B4.2.3 Bioaccumulation and biomagnification
Rapid, significant absorption of 14C-AOS by the gills of
goldfish (Carassius auratus) was seen after exposure to AOS at a
concentration of 10 mg/litre. The concentration of AOS in the gills
increased from 0.3 mg/kg after 0.5 h of exposure to 48.3 mg/kg after
3 h. AOS were not detected in the alimentary canal (Tomiyama, 1975).
Three hours is a relatively short exposure, and the authors did not
determine whether a steady state of adsorption had been achieved.
Tomiyama (1978) reported that AOS accumulated to the greatest extent
in the gills of exposed fish, with additional accumulation in the
gall-bladder. Only limited conclusions can be drawn from this study,
however, owing to the short exposure period.
B4.3 Interaction with other physical, chemical, and biological
factors
No information was available.
B5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE
Few data are available on environmental concentrations of AOS
because of the lack of an accepted analytical method for this
purpose. A modified analytical method based on MBAS-HPLC measurement
has been used to measure AOS (Takita & Oba, 1985). The concentration
in the Tama River, Japan, was calculated to be < 0.0016-0.002
mg/litre.
The annual average concentration of AOS in wastewater was
0.160-0.164 mg/litre on the basis of total MBAS concentrations of
8.4 and 8.2 mg/litre. AOS was not detected in the effluent from a
treatment plant outfall (Oba et al., 1976).
AOS can be expected to mineralize rapidly in all environmental
compartments and to be removed to a large extent during sewage
treatment. Environmental concentrations in receiving surface waters,
sediments, soils, estuaries, and the marine environment can also be
expected to be low.
B6. KINETICS
Section summary
AOS administered orally are readily absorbed by the
gastrointestinal tract of rats and are distributed throughout the
body; they are eliminated primarily in the urine and, to a lesser
extent, in the faeces within 24 h of administration. AOS applied
dermally are absorbed only minimally by intact skin. Several
metabolites have been isolated, but their chemical structures have
not been identified.
B6.1 Absorption, distribution, and excretion
14C-AOS were synthesized by sulfonation and hydrolysis of
tetradecene-1-14C. The labelled compound was composed of a
mixture of about 55% sodium 3-hydroxyalkane sulfonate
[C11H23CH(OH)-CH2SO3Na] and about 45% sodium 2-14C alkenyl sulfonate
[C11H23CH2CH214CH2SO3Na]. After oral administration of
100 mg/kg 14C-AOS (50 µCi/kg) in water to rats, the level of
radiolabel in blood reached a peak at 3 h (0.08% of the dose/ml) and
then rapidly decreased, since little radioactivity was detected 24 h
after the administration. At 4 h after administration, 0.45% of the
dose per gram of tissue was detected in liver and 0.65% in kidney,
but the levels in tissues other than the gastrointestinal tract were
< 0.1%. Thereafter, the radiolabel in organs and tissues decreased
rapidly, and 24 h after administration, about 0.8% was detected in
the caecal contents and < 0.02% in other tissues. No specific
accumulation was observed in any tissue. Within 24 h of
administration, 72% of the dose was excreted in urine and 22% in
faeces. At the end of the experiment, after four days, no 14C
residue (< 0.1% of the dose) was detected in urine or faeces.
Cumulative excretion in the bile within 12 h after administration
was about 4.3% of the radioactivity administered (Inoue et al.,
1982).
The biological half-lives of AOS and their metabolites in blood
after intravenous administration of 10 mg/kg 14C-AOS in rats were
15 and 1 h, respectively. The marked difference in half-life can be
accounted for by the fact that the binding of AOS to plasma
proteins, especially serum albumin, increased in proportion to its
concentration while that of the metabolites did not increase to any
appreciable extent. The volume of distribution of AOS was
8 litres/kg, and that of the metabolites was 0.5 litres/kg (Inoue et
al., 1982).
A dose of 0.5 ml of a 0.2% aqueous solution of 14C-AOS was
applied to the dorsal skin (4 × 3 cm) of rats with bile-duct and
bladder cannulae. The total amount absorbed through the skin was
estimated to be about 0.6% on the basis of the recoveries of 14C
in urine, bile, and the main organs over 24 h. At that time, the
level of radiolabel was higher in the liver (0.123% of dose) than in
the kidney (0.059%), spleen (0.004%), brain (0.01%), or lung
(0.012%). A total of about 0.24% of the applied dose was recovered
in these organs. After 24 h, 0.33% of the radiolabel was excreted in
the urine and 0.08% in the bile. When the solution was painted on
skin damaged by 20 applications of cellophane adhesive tape to
remove the stratum corneum, the rates of excretion were 36.3% in the
urine and 1.8% in the bile (Minegishi et al., 1977).
B6.2 Biotransformation
AOS and its metabolites were investigated in tissues and
excrement after oral administration of 100 mg/kg 14C-AOS to rats.
AOS and a metabolite more polar than AOS were detected in blood,
liver, kidney, bile, and urine by thin-layer chromatography. As most
of the 14C-labelled compounds in urine were alcoholic,
unsaturated, and of sulfonic functionality, the metabolite may be a
hydroxylated or polyhydroxylated sulfonic acid with a shorter chain
than AOS, although the precise chemical structure remains to be
elucidated (Inoue et al., 1982).
B7. EFFECTS ON LABORATORY MAMMALS AND IN VITRO TEST SYSTEMS
Section summary
The oral LD50 for AOS sodium salt in mice was 3000 mg/kg. AOS
are skin and eye irritants. Data from studies in experimental
animals are limited, but no effects were observed in a long-term
study in which oral doses of 250 mg/kg body weight per day were
administered to rats. Fetotoxicity was observed in the progeny of
rabbits administered a maternally toxic dose of 300 mg/kg body
weight per day.
The available long-term studies are inadequate to evaluate the
carcinogenic potential of AOS in experimental animals; however, in
the limited studies available in which animals were administered AOS
orally or on the skin, there was no evidence of carcinogenicity.
The limited data available also indicate that AOS are not
genotoxic in vivo or in vitro.
B7.1 Single exposures
The LD50 values for AOS (sodium salt of sulfonated C15-C18
n-olefin) in male ddy mice were 3000 mg/kg body weight by oral
administration, 1660 mg/kg by subcutaneous injection, 170 mg/kg by
intraperitoneal injection, and 90 mg/kg by intravenous injection.
The toxic effects seen at high oral doses were reduced voluntary
activity, diarrhoea, anaemia, dyspnoea, and respiratory collapse.
Clonic convulsions followed by respiratory collapse were seen in
animals given the material intravenously (Oba et al., 1968a).
B7.2 Short-term exposure
No data were available.
B7.3 Long-term exposure; carcinogenicity
B7.3.1 Mouse
The skin of Swiss-Webster mice was painted with 20% C14-C18
AOS, 25% C14-C18 AOS, 20% C14-C16 AOS, 25% C14-C16 AOS,
6.7 or 8.3% C16 1,4-sultone, water, or acetone, or remained
untreated. Animals were treated with 0.02 ml of test material on
about 1 cm2 of exposed skin three times per week for 92 weeks.
Final necropsies were conducted when the survival of each group
reached 30% (approx. 19 months