* Effect concentrations based on measured concentrations.

Invertebrates

Painter (1992) reported ranges for EC50 for the acute toxicity of AES to daphnids between 1 and 50 mg/l. However, an EC50 of 0.37 mg/l was observed in a 21-day reproduction test with Daphnia magna (Maki 1979). Also Belanger et al. (1995) observed very low effect concentrations of AES on invertebrates as both mayfly and bivalve populations were impaired at 0.77 mg/l during an 8-week mesocosmos study (Table 3.11).

Table 3.11
Effects of AES to invertebrates.

*Effect concentrations based on measured concentrations.
** 95% confidence intervals.

Fish

The LC50 values for fish are in the range between 0.39 to 450 mg/l (Table 3.12). A LOEC value of 0.22 mg/l has been reported for a chronic life cycle test with a duration of 1 year (Maki 1979). The toxicity of AES towards fish seems to increase with increasing alkyl chain length for AES with up to 16 carbons.

Table 3.12
Effects of AES to fish.

* Effect concentrations based on measured concentrations.

Toxicokinetics and acute toxicity

AES are easily absorbed in the intestine in rats and humans after oral administration. Radiolabelled C₁₁ AE₃S and C₁₂ AE₃S were extensively metabolized in rats and most of the ¹⁴C-activity was eliminated via the urine and expired air independently of the route of administration (oral, intraperitoneal or intravenous). The main urinary metabolite from C₁₁ AE₃S is propionic acid-3-(3EO)-sulfate. For C₁₂ and C₁₆ AE₃S, the main metabolite is acetic acid-2-(3EO)-sulfate. The alkyl chain appears to be oxidized to CO₂ which is expired. The EO-chain seems to be resistant to metabolism (McDermott et al. 1975; Taylor et al. 1978). Only small amounts of non-specified AES were absorbed through the skin (Painter 1992). The LD50 values after oral administration of AES range from 1,000 – 5,000 mg/kg body weight for rats (Falbe 1986; Gloxhuber and Künstler 1992; Painter 1992) indicating a low acute toxicity.

Skin and eye irritation

AES are better tolerated on the skin than, e.g., alkyl sulfates and it is generally agreed that the irritancy of AES is lower than that of other anionic surfactants. Alkyl chain lengths of 12 carbon atoms are considered to be more irritating to the skin compared to other chain lengths (Tupker 1990; Gloxhuber and Künstler 1992). The skin irritating properties of AES normally decrease with increasing level of ethoxylation (Falbe 1986; KEMI 1990). Undiluted AES should in general be considered strongly irritating. Even at concentrations of 10% moderate to strong effects can be expected. However, only mild to slight irritation was observed when a non-specified AES was applied at 1% to the skin (SFT 1991).

Subchronic and long term toxicity

A 90-day subchronic feeding study in rats with 1% of AE₃S or AE₆S with alkyl chain lengths of C_12-14 showed only an increased liver/body weight ratio (Scailteur et al. 1986). In a chronic oral study with a duration of 2 years, doses of C12-AE₃S of 0.005 – 0.05% in the diet or drinking water had no effects on rats. The concentration of 0.5% sometimes resulted in increased kidney or liver weight (Falbe 1986; Scailteur et al. 1986; Painter 1992).

Carcinogenicity

There is no indication of increased risk of cancer after oral ingestion of AES. Carcinogenic effects were not observed after skin application (Falbe 1986; SFT 1991).

Reproductive toxicity

No evidence of reproductive and teratogenic effects was seen in a two-generation study in rats fed with a mixture (55:45) of AES and linear alkylbenzene sulfonates. Dietary levels of 0.1, 0.5, and 1% were administered to the rats either continuously or during the period of major organogenesis during six pregnancies. No changes in reproductive or embryogenic parametres were observed (Nolen et al. 1975).

Classification

AES are generally classified according to Comité Européen des Agents de Surface et leurs Intermédiaires organiques (CESIO) as Irritant (Xi) with the risk phrases R38 (Irritating to skin) and R36 (Irritaing to eyes). An exception has been made for AES (2-3E0) in a concentration of 70-75% where R36 is substituted with R41 (Risk of serious damage to eyes).

AES are not included in Annex 1 of the list of dangerous substances of Council Directive 67/548/EEC.

3.3 Linear alkylbenzene sulfonates

Linear alkylbenzene sulfonates (LAS) are, by volume, the most important group of synthetic anionic surfactant today. LAS are mainly used in laundry detergents and cleaning agents. LAS are frequently used as the sodium salts as the sole surfactant in a formulation or in conjunction with other anionic, nonionic or cationic surfactants. LAS consist of an alkyl chain attached to a benzene ring in the para position to the sulfonate group. Sometimes toluene, xylene and naphthalene are used in place of benzene. The homologue distribution in commercial products covers alkyl chain lengths from C₁₀ to C₁₃ with an average chain length of C_11.6. LAS raw materials are derived from linear alkyl benzenes in which the ring is attached to a C-atom which is itself attached to two other C-atoms. The benzene ring may be attached to any of the C atoms from C₂ to C₆ but not to C₁. Structures in which the benzene ring may be attached to different C atoms are described as isomers. E.g., the structure with a C₁₂ alkyl chain and the benzene ring attached at the second alkyl carbon is designated as the C₁₂-2-isomer and abbreviated C₁₂-2.

LAS have the following structure:

3.3.1 Occurrence in the environment

The concentrations of LAS have been monitored in several environmental compartments. The monitoring conducted in the Netherlands showed that the concentrations of LAS in the effluent of seven representative municipal sewage treatment plants varied between 0.019 and 0.071 mg/l with an average value of 0.039 mg/l (Matthijs et al. 1999). Concentrations of LAS in sewage sludges have been measured in the range of 2 to 12 g/kg for primary and anaerobically digested sludge (most in the range 4-10 g/kg), whereas aerobically digested sludge and activated sludge contained 2.1-4.3 g/kg and 0.09-0.86 g/kg, respectively (Painter 1992). A monitoring of contaminants in sludge samples from municipal sewage treatment plants in Denmark showed that the concentrations of LAS varied between 0.01 and 16 g/kg (Tørsløv et al. 1997). The median concentration of all examined sludge samples (20) was 0.53 g/kg, whereas the medians were 0.02 g/kg for 11 activated sludge samples and 0.94 g/kg for 9 samples consisting of a mixture of activated and anaerobically digested sludge (Madsen et al. 1998). LAS are found in soils that are treated with sewage sludge as a fertilizer. E.g., the concentration of LAS ranged from 2.5 to 40.3 mg/kg (median 25 mg/kg) in 7 soil samples that were collected immediately after dosing of the fields with sludge; these values fell to ‘control’ values within 21 to 122 days (Painter 1992).

Aquatic sediments may also contain LAS at mg/kg levels. E.g., the LAS concentrations were 1.5 to 10 mg/kg in 10 river sediments in Germany, whereas 25 to 174 mg/kg were found at four other sites. A Danish monitoring of contaminants in coastal marine sediments showed LAS concentrations of up to 22 mg/kg (Lillebæltssamarbejdet 1998). The highest concentration of 22 mg/kg was found in a fjord in the vicinity of the discharge of effluent from a municipal wastewater treatment plant.

3.3.2 Environmental fate

Biodegradation pathways

The initial step in the biodegradation of LAS under aerobic conditions is an w -oxidation of the terminal methyl group of the alkyl chain to form a carboxylic acid. Further degradation proceeds by a stepwise shortening of the alkyl chain by b -oxidation leaving a short-chain sulfophenyl carboxylic acid. In the presence of molecular oxygen the aromatic ring structure hydrolyses to form a dihydroxy-benzene structure which is opened before desulfonation of the formed sulfonated dicarboxylic acid. The final degradation steps have not been investigated in details but are likely to occur by general bacterial metabolic routes involving a total mineralisation and assimilation into biomass (Steber and Berger 1995). Both the initial w -oxidation and the hydroxylation of the ring structure of LAS require molecular oxygen, and they are not expected to take place under anoxic conditions (Steber and Berger 1995).

Aerobic biodegradability

Numerous data on primary and ultimate biodegradation of LAS have been reported. Primary degradation of 93-97% was measured as MBAS removal in OECD tests (Schöberl et al. 1988) and removal in wastewater treatment plants are reported to exceed 85% measured as MBAS (Steber and Berger 1995). The removal of LAS in wastewater treatment plants includes sorption to the sludge particles. For different treatment plants the sorbed amount was reported to be 3-15% with a total removal of 95-99% (Painter 1992). The ultimate biodegradation of LAS in aerobic screening tests fulfils the requirements for ready biodegradability in OECD 301 tests (Table 3.13). The degradation of LAS does not lead to an accumulation of metabolites as indicated by a 95% DOC removal in a test for detection of recalcitrant metabolites (Steber and Berger 1995).

Table 3.13
Ultimate aerobic biodegradability of LAS.

Anaerobic biodegradability

Ultimate biodegradation of LAS under anoxic conditions has not been documented, and the known mechanisms that precede the aerobic mineralization, i.e. the w -oxidation and the hydroxylation of the benzene ring, require molecular oxygen. A primary anaerobic biodegradation of LAS may occur dependent on the environmental conditions. This has, e.g., been shown in continuous stirred tank reactors with anaerobically digested sludge, operated at 37° C, in which the anaerobic transformation of C₁₂ LAS corresponded to between 20 and 25% of the initial concentration (Angelidaki et al. 2000). Another recent study by Denger and Cook (1999) showed that commercial LAS and C_12-3 LAS were desulfonated under sulphur-limited anoxic conditions. The two studies show that LAS can be attacked and transformed by bacteria in the absence of molecular oxygen which implies that LAS is possibly not entirely persistent in anoxic environments. However, it is too early to assess the environmental relevance of the observed anaerobic transformation of LAS, and studies of the fate of LAS in aquatic sediments that are adapted via the continuous discharge of treated effluents should be conducted in the future. Sulphur-limited conditions are not expected to exist in anoxic sediments and, especially, marine sediments usually contain high levels of sulfate. E.g., the molar concentrations of SO₄^-- in coastal Danish sediments during summer have been measured to approximately 16 mM (Randers Fjord and Kysing Fjord) and 25 mM (Limfjorden) at the surface and approximately 5 mM in deeper layers (Jørgensen 1977; Sørensen et al. 1979).

Fate of LAS in sludgeamended soil

Since LAS are generally not degraded under anoxic conditions, levels of LAS in the g/kg range can be found in sludge which is applied to agricultural soil. The LAS in the sludge will normally biodegrade rapidly in well-aerated and aerobic soils. An extensive environmental monitoring of LAS concentrations in agricultural fields following sludge-amendment showed losses of LAS ł 98% in the majority of the sites with calculated half-lives for LAS in soil between 7 and 22 days (Holt et al. 1989; Waters et al. 1989). The field monitoring data are in agreement with laboratory studies of the mineralization of ¹⁴C-labelled C₁₂-LAS in mixtures of sludge and soil. In these studies, more than 68% of the added ¹⁴C-LAS was mineralized during 2 months, when aerobic conditions prevailed, while a lower mineralization was seen in mixtures that were partly anoxic (Gejlsbjerg et al., in press). The fate of LAS was recently evaluated for two catchment areas reflecting the eastern and western parts of Denmark (Madsen et al. 1999). The model simulations indicated that worst-case LAS concentrations in the upper 0-15 cm will be between 1 and 10 mg/kg with higher concentrations immediately after sludge application (sludge dosage: 2,000 kg/ha/year; LAS concentration 16 g/kg). The LAS concentrations in sludge are usually much slower than 16 g/kg and, hence, typical LAS concentrations in the upper 0-15 cm were estimated to between 0.1 and 1 mg/kg with higher concentrations immediately after sludge application. A substantial fraction of 98-99% of the sludge-bound LAS was predicted to degrade in the upper soil layer within one year, and the degree of leaching of LAS to depths below 1 m was predicted to be < 1.3% of the LAS applied with sludge (Madsen et al. 1999).

Bioaccumulation

Earlier studies of the bioaccumulation of LAS in aquatic organisms have mainly been performed with ¹⁴C- or ³⁵S-radiolabelled LAS. By using radioactivity measurements, whole fish BCF values for C₁₂ LAS have been determined to 108-280 for bluegill sunfish (Lepomis macrochirus; Bishop and Maki 1980), 173-245 for fathead minnow (Pimephales promelas; Kimerle et al. 1975), and 231 for zebra fish (Brachydanio rerio; Coenen 1988). Several studies show that LAS are transformed in fish (Comotto et al. 1979; Kikuchi et al. 1980; Newsome et al. 1995), but the experimental data do not allow a quantitative analysis of intact LAS and biotransformation products. Identification of metabolites suggests that biotransformation of LAS occurs via w -oxidation followed by b -oxidation. These processes lead to the formation of short-chained intermediates that are less toxic because of their lower lipophilicity compared to LAS (Newsome et al. 1995). Bioconcentration experiments using radiolabelled compounds are likely to overestimate the BCF for the intact surfactant because the radiotracer technique does not distinguish between the parent compound and radiolabelled metabolites.

Tolls (1998) examined the uptake and depuration of different LAS homologues by chemical analyses of the parent compound. The BCF tended to increase with increasing alkyl chain length but also the position of the aryl sulfonate moiety was important. A higher BCF was seen for LAS isomers with the aryl sulfonate attached to the second carbon at the alkyl chain, i.e. C₁₁-2, C₁₂-2, and C₁₃-2 (Table 3.14). As it can be seen from the data in Table 3.14, the obtained BCF values differ markedly which indicates that inter-experimental difference exists. The only difference between the experiments with C₁₂-2 LAS is the concentration of the compound in the tested mixture. Apparently, the BCF is inversely related to the concentration of the compound in the mixture, i.e. a higher BCF is obtained with decreasing test concentrations. As the toxicity of LAS is expected to decrease after the primary biotransformation, the BCF studies using chemical analyses of intact surfactant are of higher value than experiments based on radiolabelled compounds. The data in Table 3.14 indicate that the homologues in commercial LAS (i.e., C₁₀-x - C₁₃-x) have a low-to-moderate bioaccumulation potential with the exception of the C₁₃-2 LAS.

Table 3.14
Whole body BCF values in fathead minnow (Pimephales promelas). Data from Tolls (1998).

Note: C_n-in represents the ‘inner isomers’, i.e. the sum of the 3-, 4-, 5-, 6-, and 7-sulfophenylalkanes, in contrast to the 2-isomer. The values in parentheses specify the relative standard variation in %.

3.3.3 Effects on the aquatic environment

Numerous studies have been performed to determine the effects of LAS towards aquatic organisms. The aquatic effect concentrations that were observed in these studies are highly variable. This variation is partly related to the testing of different isomers and homologues, but it may also be due to the specific test conditions and species. The length of the alkyl chain is an important factor determining the aquatic toxicity. In general, the homologues with the highest number of carbons in the alkyl chain are more toxic than are those with shorter alkyl chains. Today, commercial LAS have a homologue distribution between C₁₀ and C₁₃ with a typical average alkyl chain length of C_11.6.

Algae

The widest range in the toxicity of LAS towards species belonging to the same group is found for algae (Table 3.15). Approximately 90% of the data found in the literature fall between 0.1 and 100 mg/l. Typical ranges of EC50 values are 1 to 100 mg/l for fresh water species and < 1 to 10 mg/l for marine species (Painter 1992). A very low EC100 value of 0.025 mg/l was determined for Gymnodium breve (Hitchcoch and Martin 1977). Previous studies in which Gymnodium breve was exposed with AES confirm that this species is highly sensitive to surfactants (Kutt and Martin 1974), and occasionally available data for Gymnodium breve should therefore not be used for comparison of the aquatic toxicity between various surfactants.

Table 3.15
Effects of LAS to algae.

* Effect concentrations based on measured concentrations.
** Parentheses indicate 95% confidence intervals.

Invertebrates

LC50 values have been found in the range of 1 to 10 mg/l when Daphnia magna were exposed with LAS homologues between C₁₀ and C₁₃. The acute toxicity of LAS to Daphnia magna generally increases with increasing alkyl chain length. This is illustrated by studies performed by Maki and Bishop (1979) showing that LAS homologues ł C₁₄ produce EC50 values below 1 mg/l (Table 3.16). Similar results were obtained in a study of LAS homologues between C₁₀ to C₁₄ as the 48 h-LC50 values were 1.2 mg/l for C₁₄ LAS and 53.1 mg/l for C₁₀ LAS (Kimerle and Swisher 1977). A study with the marine crustacean Acartia tonsa indicated that a C_10-13 LAS affected the survival at 0.54 mg/l (LC50) and the development rate at 0.51 mg/l (EC50) after 8 days of exposure. The 48 h-LC50 that was obtained in the same study with Acartia tonsa was 2.1 mg/l (Kusk and Petersen 1997). Metabolites from biotransformation of LAS are reported to have a much lower toxicity to invertebrates compared to the toxicity of the intact surfactant (Painter 1992).

Table 3.16
Effects of LAS to Daphnia magna unless otherwise indicated.

* Effect concentrations based on measured concentrations.
** 95% confidence intervals.

Fish

The toxicity of LAS to fish generally increases with increasing alkyl chain length, and approximately a 10-fold difference in toxicity between homologues separated by two carbon atoms has been observed. As also noted for invertebrates, fish are less susceptible to metabolites from biotransformation of LAS (Painter 1992). LC50 values below 1 mg/l were found for C_11.9 (0.71 mg/l), C₁₃ and C₁₄ (both 0.4 mg/l) in studies with fathead minnow (Table 3.17) and for C_10-15 (0.36 mg/l; 96 h) in a study with rainbow trout (Brown et al. 1978).

Table 3.17
Effects of LAS to fathead minnow (Pimephales promelas) unless otherwise indicated.

* Effect concentrations based on measured concentrations.
** 95% confidence intervals.

Sediment organisms

LAS sorb to sediment with partition coefficients of 50 to 1,000. The toxicity of LAS bound to sediment is relatively low compared to LAS in solution. NOEC and LOEC values were as high as 319 and 993 mg LAS/kg, respectively, for the sediment-living Chironomus riparius. The corresponding NOEC for LAS in solution was as low as 2.4 mg/l indicating that only a small fraction of the sorbed LAS was bioavailable (Painter 1992). Bressan et al. (1989) investigated the effects of LAS dissolved in water and found acute effects in the range of 0.25 to 200 mg/l dependent of the species. Copepods and embryos of the sea urchin Paracentrotus lividus were the most sensitive organisms. LAS dissolved in water may also cause chronic effects like reduction of the growth rate of the marine mussel Mytilus galloprovincialis. LAS sorbed to sediments did not have similar effects. No alterations of the treated organisms were observed although the LAS concentrations in the sediment were 3 to 10 times higher than the effect concentrations observed for LAS in water. The 96 h-LC50 values for sediment-bound LAS were 182.5 mg/kg and 200 mg/kg for the bivalve molluscs Unio elongatulus and Anodonata cygnea.

3.3.4 Effects on human health

Toxicokinetics and acute toxicity

LAS are readily absorbed by the gastrointestinal tract after oral administration in animals. LAS are not readily absorbed through the skin (IPCS 1996). The bulk is metabolized in the liver to sulfophenylic carboxyl acids. The metabolites are excreted primarily via the urine and faeces. The main urinary metabolites in rats are sulfophenyl butanoic acid and sulfophenyl pentanoic acid. Accumulation of LAS or its main metabolites has not been established in any organ after repeated oral ingestion (SFT 1991).

Dodecylbenzene sulfonate (C₁₂ LAS) was administrated daily in the diet at a dose level of 1.4 mg/kg body weight for 5 weeks. Of the administered dose of C₁₂ LAS, 52.4% was excreted in faeces and 29.4% in urine during the dosing period. A single application of 0.385 mg C₁₂ LAS per rat resulted in a total elimination of 85% within the first 24 hours and 95% within 10 days (Lay et al. 1983). No data on skin absorption were identified, but the skin absorption of anionic surfactants is generally considered to be very low.

Table 3.18
Acute toxicity (LD50) of alkyl benzene sulfonates.

No serious injuries or fatalities in man have been reported following accidental ingestion of LAS-containing detergent (Painter 1992; IPCS 1996). The main clinical signs observed after oral administration to rats of doses near or greater than the LD50 values consisted of reduced voluntary activity, diarrhoea, weakness etc. Death usually occurred within 24 hours of administration. Rats appear to be more sensitive to LAS than mice (IPCS 1996).

Skin and eye irritation

LAS and branched alkylbenzene sulfonates may cause irritation of the eyes, skin and mucous membranes. LAS are relatively more irritating to the skin than the corresponding branched alkylbenzene sulfonates (KEMI 1990). The potential of LAS to irritate the skin depends on the concentration applied. LAS have been classified as irritating to skin at concentrations above 20% according to EU-criteria. Human skin can tolerate contact with solution of up to 1% LAS for 24 hours resulting in only mild irritation (IPCS 1996). Application of > 5% LAS to the eyes of rabbits produced irritation. Concentration of < 0.1% LAS produced mild to no irritation (CIRP 1993).

Sensitization

Skin sensitization was not seen in 2,294 volunteers exposed to LAS or in 17,887 exposed to formulations of LAS (Nusair et al. 1988).

Subchronic and long-term toxicity

A feeding study indicated that LAS, when administered for 2 years at extremely high levels (0.5%) in the diets to rats, produced no adverse effects on growth, health or feed efficiency (Buehler et al. 1971).

Mutagenicity and carcinogenicity

The mutagenic potential of LAS was tested using Salmonella typhimurium strains, using Ames test. In these studies, LAS was not mutagenic (Inoue and Sunakawa 1980). The available long-term studies are inadequate for evaluating the carcinogenic potential of LAS in laboratory animals. The studies available (oral administration to rats and mice) do not show any evidence of carcinogenicity (Gloxhuber and Künstler 1992; IPCS 1996).

Reproductive toxicity

LAS was applied daily from day 0 through day 20 of gestation to the shaved skin of pregnant rats. The applied concentrations of LAS were 0.05-0.5%, and the doses remained on the skin. Furthermore, concentrations of 1%, 5% and 20% LAS were applied to the skin of pregnant rats, and these doses were removed 30 minutes after exposure. The only effects attributed to LAS were reduced body weight in the dams given the highest concentration (20%), and local skin changes in the dams which received the two highest concentrations (5% and 20%). There were no findings indicative of effects of LAS on the foetal parameters evaluated and no indication of teratogenic or embryotoxic effects (Daly and Schroeder 1980).

A 20% LAS solution (0.1 ml) was applied twice daily to the dorsal skin (2 x 3 cm) of pregnant 1CR/Jc1 mice during the preimplantation period (days 0-3 of gestation). A significant number of embryos collected on day 3 were severely deformed or remained at the morula stage. Nomura et al. (1980, 1987) reported that the number of embryos in the oviducts was significantly greater for the mice dosed with LAS as compared to the control mice used in that study. No pathological changes were detected in the major organs of the dams.

In general no specific effect of LAS on reproductive processes has been seen, although dosages causing maternal toxicity may also induce some effects on reproduction. No teratogenic effects attributed to LAS exposure were observed (Gloxhuber and Künstler 1992; IPCS 1996).

Classification

LAS are classified as Irritant (Xi) with the risk phrases R38 (Irritating to skin) and R41 (Risk of serious damage to eyes) according to CESIO (CESIO 2000).

LAS are not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC.

3.4 Secondary alkane sulfonates

Alkane sulfonates are used in liquid detergents like, e.g., dishwashing agents, cleaning agents, and hair shampoos, frequently in combination with AES. Commercial products are almost exclusively composed of secondary alkane sulfonates (SAS) with the following structure:

The alkyl chain (R₁+R₂), normally C₁₁ to C₁₇, is linear and the SO₃^- group is placed randomly along the alkyl chain. Thus, e.g., C₁₄ alkane sulfonate is a mixture of the six isomers 2-, 3-, 4-, 5-, 6- and 7-sulfotetradecane.

No data were found on the occurrence of SAS in the environment.

3.4.1 Environmental fate

Biodegradation pathways

The  of SAS biodegradation has only been scarcely investigated. A pathway similar to that of LAS involving an initial w /b -oxidation is an obvious assumption, but this has not yet been proven. One study suggests that the initial degradation step is a desulfonation requiring molecular oxygen (Painter 1992). This process involves the formation of a hydroxysulfonate which is hydrolysed to inorganic sulfate and a ketone. The ketone is subsequently oxidised to an ester which is cleaved to acetic acid and the corresponding alcohol (Painter 1992).

Aerobic biodegradability

SAS undergo rapid primary biodegradation with MBAS removals higher than 90% within a few days (Swisher 1987). Removals of 96% were seen in the OECD screening test for primary biodegradation (Schöberl et al. 1988). In activated sludge simulation tests, 96% of C_10-18 SAS was removed, while the parent C_13-18 SAS was removed by 83-96% (Painter 1992). The fate of a ¹⁴C-labelled C₁₇ SAS was followed in a continuous activated sludge test to illustrate the ultimate biodegradation. After 3 days, 47% of the added C₁₇ SAS were detected as ¹⁴CO₂ and 25% were incorporated into sludge biomass (Steber and Berger 1995). The ultimate biodegradability of SAS fulfils the criteria for ready biodegradability in OECD 301 tests (Table 3.19).

Table 3.19
Ultimate aerobic biodegradability of SAS.

Anaerobic biodegradability

Wagener and Schink (1987) investigated the anaerobic biodegradability of SAS in tests incubated with either digested sludge or creek sludge and came to the conclusion that alkyl sulfonates are not degraded under anoxic conditions.

Bioaccumulation

No experimental data describing the bioaccumulation potential of SAS were found in the literature.

3.4.2 Effects on the aquatic environment

Algae

The toxicity of various SAS homologues was determined in tests with Chlamydomonas variabilis. After 24 hours of exposure at 20° C, there was a tendency to an increased toxicity with increasing chain length. The EC50 values were 125 mg/l for C_10.3, 74.9 mg/l for C_11.2, 32.4 mg/l for C₁₄, 15.8 mg/l for C₁₅, 9.42 mg/l for C₁₆, 3.93 mg/l for C₁₇, 3.71 mg/l for C_18.9, and 8.47 mg/l for C_20.7 (Lundahl and Carbridenc 1978).

Invertebrates

The same tendency to an increased toxicity of SAS with increasing chain length was seen in tests with Daphnia magna. The tests with Daphnia magna showed 24 h-EC50 values at 319 mg/l for C_10.3, 133 mg/l for C_11.2, 111 mg/l for C₁₄, 34.2 mg/l for C₁₅, 30.1 mg/l for C₁₆, 12.3 mg/l for C₁₇, 3.31 mg/l for C_18.9, and 6.30 mg/l for C_20.7 (Lundahl and Carbridenc 1978). Schöberl et al. (1988) reported an EC50 range of 8.7-13.5 mg/l for daphnia in studies with C_13-18 SAS, whereas Painter (1992) reported a lower EC50 range of 0.7-6 mg/l for C_15-18 SAS.

Fish

Also for fish the longer chain length SAS are more toxic than the shorter chained homologues (Table 3.20). This has been shown both for minnow (Lundahl and Carbridenc 1978) and for bluegill sunfish (Painter 1992). Schöberl et al. (1988) reported a range of LC50 values of 3-24 mg/l for C_13-18 SAS in tests with fish species that were not specified.

Table 3.20
Effects of SAS to fish.

3.4.3 Effects on human health

Toxicokinetics and acute toxicity

SAS are readily absorbed from the gastrointestinal tract of rats after oral administration. Following administration of C₁₂ and C₁₆ SAS the main metabolite is butyric acid-4-sulfonate. This metabolite is eliminated in the urine (McDermott et al. 1975; Taylor et al. 1978).

The acute toxicity of a SAS of non-specified chain length in the rat was moderate with LD50 values between 1,000 and 3,000 mg/kg body weight when administrated by the oral route (Falbe 1986; SFT 1991; Gloxhuber and Künstler 1992).

Skin and eye irritation

The irritating potential of SAS to skin is almost the same as that of alkyl sulfates. Concentrations of more than 20% alkane sulfonate are strongly irritating to the skin of rabbits (SFT 1991).

Chronic toxicity, carcinogenicity, mutagenicity

Subchronic studies with rats receiving 300 mg SAS/kg body weight/day orally for 45 and 90 days revealed no adverse effects. Similarly, rats fed 0.5% SAS in their diets for 91 days developed no adverse symptoms (Scailteur et al. 1986; Painter 1992). There was no indication of increased risk of cancer after oral ingestion of SAS in studies that were not further specified (Falbe 1986; SFT 1991).

Classification

The skin irritating potential of SAS is about the same as for alkyl sulfates. SAS may therefore also be classified as Irritant (Xi) with R38 (Irritating to skin) and R41 (Risk of serious damage to eyes).

SAS are not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC.

3.5 a -Olefine sulfonates (AOS)

a -Olefine sulfonates (AOS) are used in laundry powder detergents, liquid dishwashing agents, as well as in hair shampoos, and mainly in Japan and the USA (Steber and Berger 1995). AOS consist of a mixture of alkene sulfonates (about 60-65%) and hydroxyalkane sulfonates (about 30-40%). The normally linear C-chain in alkene 1-sulfonates and hydroxyalkane 1-sulfonates may contain 11 to 20 carbons with 14 to 18 carbons as the usual range (Painter 1992).

The alkene sulfonates have the structure:

H₃C-(CH₂)_m-CH=CH-(CH₂)_n-SO₃^-Na⁺