Canadian Environmental Protection Act, 1999
Federal Environmental Quality Guidelines
(PDF Format - 181 KB)
Table of Contents
- Substance Identity
- Fate, behaviour and partitioning in the environment
- Ambient concentrations
- Mode of action
- Freshwater toxicity
- Federal Water Quality Guideline Derivation
- Considerations in Guideline Implementation
- List of Acronyms
Federal Environmental Quality Guidelines (FEQGs) provide benchmarks for the quality of the ambient environment. Where the FEQG is met there is low likelihood of adverse effects on the protected use (e.g., aquatic life or the wildlife that may consume them). They are based on the toxicological effects or hazards of specific substances or groups of substances and do not take into account analytical capability or socio-economic factors. FEQGs serve three functions: first, they can be an aid to prevent pollution by providing targets for acceptable environmental quality; second, they can assist in evaluating the significance of concentrations of chemical substances currently found in the environment (monitoring of water, sediment, and biological tissue); and third, they can serve as performance measures of the success of risk management activities. The use of FEQGs is voluntary unless prescribed in permits or other regulatory tools. Thus FEQGs, which apply to the ambient environment, are not effluent limits or “never-to-be-exceeded” values but may be used to derive effluent limits. The development of FEQGs is the responsibility of the Federal Minister of Environment under the Canadian Environmental Protection Act, 1999. The intent is to develop FEQGs as an adjunct to risk assessment/risk management of priority chemicals identified in the Chemicals Management Plan (CMP) or other federal initiatives. This factsheet describes the Federal Water Quality Guidelines (FWQGs) for the protection of aquatic life for alcohol ethoxylates and some related substances (see Figure 1) and is based on information obtained to 2004. No FEQGs have been developed for the soil, sediment or biological tissue compartments at this time.
Federal Water Quality Guideline: 70 µg/L
(for representative alcohol ethoxylate homologue C13.7 EO5)
Figure 1. Species sensitivity distribution (SSD) for freshwater toxicity data for the representative alcohol ethoxylate homologue C13.7 EO5 and associated effect levels for freshwater life. Guidelines for individual homologues are given in Table 3
Alcohol ethoxylates (AEs) are part of the alcohol alkoxylates class which also includes alcohol propoxylates and butoxylates. Alcohol ethoxylates are a class of nonionic surfactants that contain a hydrophobic alkyl chain attached via an ether linkage to a hydrophilic ethylene oxide (EO) chain and have the general structure R(OCH2CH2)nOH. The alkyl chain, R, can vary in length and in the degree of linearity, but is typically between 8 and 18 carbons long (for detergent range surfactants). The EO chain can also vary in length from 1 to 40 EO units. An AE with the structure C9-11EO6.5, for example, contains a range of alkyl chain lengths of 9-11 and averages 6.5 EO units per alkyl chain. It should be noted that, although this is the general description of the mixture, other homologues are present. For example, an AE produced from a C12-15 alcohol mixture can yield up to 100 different surfactants (Raney 2000). Due to the large number of possible AEs, many different CAS numbers and trade names exist for AEs (Table 1). Because of their ecological and/or human health concerns, many of these substances were identified as priorities for further action under the CMP.
The hundreds of different possible AEs each have slightly different chemical and physical properties (Hennes-Morgan and De Oude 1994), however, the presence of a strong hydrophilic (ethoxylate chain) and strong hydrophobic (alkyl chain) moiety linked together gives them their characteristic surfactant properties (Swisher 1987). AEs concentrate at surfaces and interfaces in aqueous solutions and create a surface film which reduces the surface tension of water and alters the wetting properties between water and solids (Aveyard 1984; Swisher 1987). The solubility of AEs in water results from the presence of the hydrophilic group (Aveyard 1984).
Alcohol ethoxylates are currently some of the most commonly manufactured and utilised nonionic surfactants in Canada and world-wide (Camford Information Services 1997; Campbell 2002). AEs are predominantly used in consumer and industrial products such as laundry detergents and all-purpose cleaners, and are used to a lesser extent by agriculture, cosmetic, textile, paper, and oil sectors (Talmage 1994; Camford Information Services 1997; Madsen et al. 2001). Total production capacity of AEs in Canada in 1996 was estimated at 72 000 t (Camford Information Services 1997). In 2000, a total of 216 800 t of AEs were consumed in Canada and the United States (Modler et al. 2002). Of this total, approximately 41% was used in laundry liquids, 20% in laundry powders, 3% in dishwashing liquids, 9% in other household cleaners, and the remaining 27% was consumed through other uses. In addition, another 411 408 t of AEs were consumed in Canada and the United States in 2000 in the production of alcohol ethoxysulfates, a group of anionic surfactants. In 2002, it is estimated that over 500 000 t of fatty alcohols and fatty-alcohol-based surfactants were consumed in North America (Modler et al. 2004). Of this, approximately 67% consisted of AEs or sulfated AE.
Table 1. Chemicals to which Federal Water Quality Guidelines for alcohol ethoxylates apply
|CAS Registry Number||Chemical Name|
|61791-28-4||Alcohols, tallow, ethoxylated|
|66455-14-9||Alcohols, C12-13, ethoxylated|
|68002-96-0||Alcohols, C16-18, ethoxylated propoxylated|
|66455-15-0||Alcohols, C10-14, ethoxylated|
|68002-97-1||Alcohols, C10-16, ethoxylated|
|68131-39-5||Alcohols, C12-15, ethoxylated|
|68154-96-1||Alcohols, C14-18, ethoxylated|
|68154-98-3||Alcohols, C14-18, ethoxylated propoxylated|
|68155-01-1||Alcohols, C16 and C18-unsatd., ethoxylated|
|68213-23-0||Alcohols, C12-18, ethoxylated|
|68213-24-1||Alcohols, C12-16, ethoxylated propoxylated|
|68439-45-2||Alcohols, C6-12, ethoxylated|
|68439-46-3||Alcohols, C9-11, ethoxylated|
|68439-50-9||Alcohols, C12-14, ethoxylated|
|68439-51-0||Alcohols, C12-14, ethoxylated propoxylated|
|68439-54-3||Alcohols, C11-13-branched, ethoxylated|
|68526-94-3||Alcohols, C12-20, ethoxylated|
|68551-12-2||Alcohols, C12-16, ethoxylated|
|68551-13-3||Alcohols, C12-15, ethoxylated propoxylated|
|68551-14-4||Alcohols, C11-15-secondary, ethoxylated propoxylated|
|68920-66-1||Alcohols, C16-18 and C18-unsatd., ethoxylated|
|68951-67-7||Alcohols, C14-15, ethoxylated|
|68991-48-0||Alcohols, C7-21, ethoxylated|
|69013-19-0||Alcohols, C8-22, ethoxylated|
|69227-20-9||Alcohols, C16-22, ethoxylated|
|69227-21-0||Alcohols, C12-18, ethoxylated propoxylated|
|70879-83-3||Alcohols, C6-10, ethoxylated|
|71243-46-4||Alcohols, C8-16, ethoxylated|
|73049-34-0||Alcohols, C16-20, ethoxylated propoxylated|
|106232-83-1||Alcohols, C12-15-branched and linear, ethoxylated|
|111905-53-4||Alcohols, C13-15-branched and linear, butoxylated ethoxylated|
|111905-54-5||Alcohols, C13-15-branched and linear, ethoxylated propoxylated|
|139626-71-4||Alcohols, C14-100, ethoxylated|
Alcohol ethoxylates are synthetic chemicals that do not occur naturally. Since the majority of AEs are used in cleaners and detergents, the largest recipient of chemical inputs is the aquatic environment, primarily through wastewater effluent as a result of consumer disposal practices (Holt et al. 1992). Other potential sources of exposure exist within the industrial sectors where AEs are manufactured or used (industrial cleaners, pulp and paper, chemical manufacture). Natural degradation processes and wastewater treatment techniques remove a large proportion of AE from water, however there is still potential for aquatic exposure. Due to the hydrophobic moiety of AE compounds, once introduced into the aquatic environment they have the potential to adsorb to particulate matter and be deposited to sediments (McAvoy and Kerr 2001). An additional pathway for environmental exposure to AEs is via direct application to soils such as in sewage sludge, and through the use of septic systems which use the soil to treat and disperse wastewater (Nielsen et al. 2002).
Fate, behaviour and partitioning in the environment
By far the most significant fate process for AEs is aerobic microbial biodegradation. Due to the down-the-drain disposal pattern of AEs, a majority of their biodegradation occurs in the sewage system (Nielsen et al. 2002), wastewater treatment plants (WWTP) (McAvoy et al. 1998), septic systems (Matthijs et al. 1995) and to a lesser extent in natural waters (Vashon and Schwab 1982). Multibranched AEs, which make up significantly less of the commercial AE market in Canada (Campbell 2002), are slower to degrade than linear AEs (Birch 1984; Marcomini et al. 2000) but can still be considered readily degradable. Intermediate degradation products of AEs can include free fatty alcohols, polyethylene glycols (PEG), and carboxylic fatty acids.
Adsorption of AEs plays a minor role in their aquatic fate, with sorption to humic materials in water, suspended and bed sediments, and sludge in wastewater treatment facilities. AEs are not expected to volatilize to the atmosphere (Kiewiet et al. 1997).
AEs are rapidly taken up across the gills in fish (Bishop and Maki 1980; Wakabayashi et al. 1987; Tolls et al. 1994; Newsome et al. 1995) and are rapidly metabolised and eliminated from fish (Bishop and Maki 1980; Wakabayashi et al. 1987; Newsome et al. 1995). Based on the high elimination rates, Tolls et al. (2000) concluded that rapid biotransformation was taking place in fathead minnows (Pimephales promelas), and that AEs were not stored in the fish. Metabolites formed during biotransformation are of less concern than the parent compound as the short-chained metabolites have low lipophilicity and therefore are less toxic (Roberts 1991; Newsome et al. 1995). Bioconcentration factors reported for AEs are typically less than 300 (e.g., Tolls et al. 2000) and would not meet the criteria for Persistence and Bioaccumulation Regulations (Government of Canada 2000). However, log octanol/water partitioning coefficients (log Kow) for various AEs have been estimated to range from approximately 3 to 7, suggesting that some AEs may tend to bioaccumulate (Müller et al. 1999b).
Any AEs detected in the environment are the result of anthropogenic releases. Free fatty alcohols, however, which are found in mixtures of AE homologues, can also occur naturally. Free fatty alcohols are generally analysed with and quantified as AEs, and may therefore result in an over estimation of AE levels. There is currently no information on ambient concentrations of AEs in the Canadian environment.
In August 2003, effluents were sampled from eight Canadian municipal WWTP (P&G – Shell 2003; Sherren et al. 2003; Eadsforth et al. 2006). The total AE concentrations for these eight plants ranged from 1.0 to 22.7 µg/L, with an average of 6.8 µg/L (Sherren et al. 2003). It is notable that the two highest concentrations occurred in effluents from the two plants that did not use activated sludge treatment. The measurements of total AEs included those species with no ethoxylate groups (i.e., free fatty alcohols). Among the eight municipal WWTP, free alcohols accounted for 21 to 62% of the total AE concentration. Although some of this free alcohol likely resulted from the degradation of AEs, some may also have originated from other sources including other alcohol-based surfactants such as alkyl sulfates and alcohol ethoxysulfates, other uses of alcohols, and as metabolic by-products from microbial degradation of animal and vegetable matter in the municipal WWTP (Scott Belanger, Procter & Gamble, pers. comm. 2004, Mudge et al. 2008). Ratios of the molar concentrations of free alcohol to AEs in the effluent were determined for each alkyl chain length. Using these ratios, it is therefore possible to determine “alcohol caps”, that is, maximum concentrations of free alcohol in effluents that could have originated from AEs. The alcohol cap is defined as the ratio (for each chain length) between the effluent alcohol that could have been derived from AE and the total ethoxylated alcohols in the effluent (EO > 1). The cap values derived from Stephenson et al. (2004) were 0.58, 0.63, 0.33, 0.25, 0.26 and 0.05 for C12, C13, C14, C15, C16 and C18, respectively. The cap is applied in any calculation only if the effluent measured alcohol concentration exceeds: cap x Sum (EO1-20). When alcohol caps are applied to the monitoring data from the eight Canadian municipal WWTP, the concentrations of total AEs ranged from 0.8 to 11.5 µg/L with an average of 4.9 µg/L (P&G –Shell 2003).
From the study of AE concentrations in Canadian municipal wastewater effluents (P&G – Shell 2003; Sherren et al. 2003), it is also possible to determine the average homologue distribution. Applying an alcohol cap, the average alkyl chain length was 13.68 carbons long, and the average ethoxylate chain consisted of 5.03 ethoxylate units (Scott Belanger, Procter & Gamble, pers. comm. 2004). In other words, the average homologue distribution in Canadian municipal wastewater effluents is C13.7EO5.
Mode of action
Although the exact mechanism by which AEs affect aquatic organisms is not fully understood, it is likely that they act through nonspecific narcosis depending on the number of EO units (Roberts 1991; Dorn et al. 1997a; Müller et al. 1999a). Narcosis is a nonspecific, reversible mode of toxic action in which the presence of hydrophobic organic chemicals causes a disruption of cellular activity. Many researchers suspect that surfactants such as AE can disrupt gill membranes in fish, invertebrates and amphibians, and cause the gill epithelial cells to swell and secrete mucous (Moore et al. 1987; Cardellini and Ometto 2001). These cell membrane disruptions can affect diffusion of oxygen across the gills, ultimately resulting in suffocation (Moore et al. 1987; Cardellini and Ometto 2001).
In algal cells, surfactants such as AE are thought to denature and bind proteins in the cell wall, thereby altering the permeability of the cell membrane to nutrients and chemicals (Lewis 1990). The susceptibility of algal species can vary depending on the thickness and chemical composition of their cell walls. Algal species with thicker cell walls will be affected less by exposure to surfactants (Nyberg 1988; Lewis 1990). Hydrophobic surfactants will more easily penetrate algal species with high lipid and protein content in their cell walls (Lewis 1990).
In some cases, AE toxicity can also result from physical surface active effects. When applied to the surface of water bodies, AE can form a film that alters the water surface tension, thereby affecting surface breathing or water striding invertebrate larvae or adults (Mulla et al. 1983).
Fish and invertebrates are generally more sensitive to AE than plants or algae (Bisop and Perry 1981; Dorn et al. 1997b). The toxicity of individual AE homologues is a function of their chemical structure. The toxicity of AE increases with increasing alkyl chain length (Wong et al. 1997; Dorn et al. 1997a; Lizotte, Jr. et al. 1999; Ghirardini et al. 2001), and with decreasing ethoxylate chain length (Macek and Krzeminski 1975; Maki and Bishop 1979; Yamane et al. 1984; Wong et al. 1997; Raney 2000). In addition, to the length of the alkyl chain, the structure of the chain can also affect its toxicity as linear AEs are more toxic than branched AEs (Dorn et al. 1993; Kaluza and Taeger 1996; Ghirardini et al. 2001). Other factors that have been demonstrated to affect toxicity are the location of attachment between the alcohol and ethoxylate chains (e.g., primary vs. secondary alcohols) (Kurata et al. 1977), the homologue distribution of the ethoxylate chain (Garcia et al. 1996), water hardness (Tovell et al. 1975), and temperature (Lewis and Hamm 1986).
Commercial AEs consist of a mixture of various AE homologues which differ in their relative toxicity. As a result, direct comparisons cannot be made among all of the available AE toxicity studies because they involve tests on numerous different mixtures. Commercial AE mixtures that are used in toxicity testing will also differ in their homologue distributions from the AE mixtures that are found in the aquatic environment following varying degrees of biodegradation. Therefore, direct comparisons cannot be made between results of laboratory toxicity tests and total AE concentrations that are measured in ambient waters. For the purpose of assessing acceptable levels of AE in water, a solution is to normalize the toxicity data to a common homologue distribution.
A monitoring survey of Canadian municipal wastewater effluents identified the average AE homologue distribution, in Canadian effluents, adjusted for an alcohol cap, to be C13.7EO5 (Scott Belanger, Procter & Gamble, pers. comm. 2004). Where the exact homologue distribution of AEs at a particular site is unknown, this average distribution can be used as the structure to which all acceptable reported AE toxicity data are normalized. It should be noted, however, that this average distribution is based solely on analyses of municipal wastewater effluents. Where AEs in ambient Canadian waters originate from other types of sources, such as discharges from industries, it is possible that the average homologue distribution will differ, due to differences in the types of parent AEs that are used.
AE toxicity data were normalised based on the predicted effect concentrations for both the test structure and the normalized structure as determined through the use of quantitative structure-activity relationship models (QSARs). A number of QSARs have been developed for AEs, both for acute and chronic exposure, as well as various types of organisms.
The general formula used to normalize AE toxicity data to a common homologue distribution is as follows:
normalized ECenv = reported ECtest × (predicted ECenv / predicted ECtest)
normalized ECenv = normalized effect concentration for environmental distribution
reported ECtest = effect concentration from a toxicity test on a commercial distribution
predicted ECenv = effect concentration from a QSAR based on environmental distribution
predicted ECtest = effect concentration from a QSAR based on a commercial distribution
In normalizing the aquatic toxicity data, corrections were also made for sorption. The sorption correction was made to account for homologue-specific sorption to effluent suspended solids, with the assumption that sorbed fractions are not bioavailable, and therefore are not contributing to the toxicity of the AE distribution. Adjustments for sorption were made using a QSAR based on a regression between Kd or Koc and chain length and ethoxylate number (van Compernolle et al. 2006). This is made possible because all sorption studies have utilized pure materials and not mixtures. The sorption corrections were used to adjust the average environmental homologue distribution. Corrections for solubility were made using a QSAR developed by Hansch et al. (1968) based on log Kow. Both the solubility and sorption corrections were used to adjust the average environmental homologue distribution. The solubility corrections, however, were found to have little effect since the AE homologue concentrations measured in Canadian municipal wastewater effluents never exceeded solubility limits. For the AEs used in the toxicity tests it was assumed that all homologues were bioavailable and below solubility limits, so no corrections were needed.
The sensitivities for AEs overlap among taxa (Table 2). The most sensitive species reported for AE is an invertebrate species bivalve (Corbicula fluminea), followed by the water flea (Daphnia magna). A mayfly (Isonychia bicolour) was the least sensitive invertebrate species in the data set. Among fish species, rainbow trout (Oncorhynchus mykiss) and bluegill (Lepomis macrochirus) were the most and least sensitive species, respectively. In general, plants tend to be less sensitive than invertebrates and fish, though some plant species may be quite sensitive (i.e., rotifers). Among plants, Navicula pelliculosa was most sensitive whereas Chlorella vulgaris was least sensitive.
Federal Water Quality Guideline Derivation
The FWQGs developed here identify benchmarks for aquatic ecosystems that are intended to protect all forms of aquatic life for indefinite exposure periods. Although AEs are rapidly degraded in water under field conditions, aquatic organisms may be chronically exposed to AEs due to their continuous release in municipal wastewater effluents. The acceptable data identified for developing the AE guidelines consist of three fish species, eight invertebrate species, and six plant species (Table 2).
Table 2. Toxicity data used for developing the Federal Water Quality Guideline for alcohol ethoxylates. Concentrations were normalised to C13.7EO5
|56d MATC |
|35d EC20 |
|7d EC20 |
|10d LC20 |
|4d LC20 |
|72h EC20 |
Notes: EC20= Concentration at which there is an adverse effect on 20% of the population; LC20 = concentration lethal to 20% of the population; MATC = maximum acceptable toxicant concentration.
The species sensitivity distribution (SSD) curve was fitted using the toxicity data (Figure 1). Each species for which appropriate toxicity data were available was ranked according to sensitivity, and its centralized position on the SSD was determined. Following the CCME protocol (CCME 2007), several cumulative distribution functions (CDFs) were fit to the data using regression methods and the best model was selected based on consideration of goodness-of-fit and model feasibility (for details see CCME 2007). The logistic model provided the best fit of the models tested and the 5th percentile (HC5) of the SSD plot is 70 µg/L,with lower and upper confidence limits of 50 and 110 µg/L, respectively (Figure 1).
The recommended FWQG is 70 µg/L. The guideline represents the concentration below which you would expect either no, or only a low likelihood of adverse effects on aquatic life. In addition to this guideline, two other concentration ranges are provided for use in risk management. At concentrations between greater than the 5th and 50th percentile of the SSD (>70-250 µg/L) there is a moderate likelihood of adverse effects to aquatic life. Concentrations greater than the 50th percentile (>250 µg/L) have a higher likelihood of causing adverse effects. Risk managers may find these additional concentration ranges useful in defining short-term or interim risk management objectives for a phased risk management plan. The moderate to higher concentration range may also be used in setting less protective objectives for waters that are already highly degraded or where there are socio-economic considerations that preclude the ability to meet the FWQG (i.e., the 5th percentile).
This FWQG represents the average homologue distribution of AEs and can be used where the average homologue distribution of AE in the ambient water is known to be C13.7EO5, or this can be used as a default guideline if the average AE homologue distribution at a site is unknown. Guideline values have also been
determined for a suite of individual AE homologues that could occur in the environment (Table 3). These were determined using the same method used for the default guideline, by normalizing the toxicity data to each individual homologue and creating a SSD distribution for each individual homologue. Where possible, it is recommended that concentrations of each individual AE homologue at a site be determined and then compared with the corresponding guideline value given in Table 3 through the use of a Toxic Unit approach.
Table 3. Federal Water Quality Guideline values (mg/L) for specific AE homologues[*] (see text for explanation)
To derive the guideline for specific homologues (each cell) shown in Table 3, the following steps were taken. First, for each homologue, the calculated log P (C log P) (a theoretical measure of hydrophobicity based on structural modelling) value was determined as a function of carbon chain length and ethoxylate length measure. The C log P was used in place of log Kow as it is difficult to obtain octanol-water partitioning coefficients (Kows) for AEs because, as surfactants, they do not partition well into either liquid, but rather tend to emulsify the octanol-water interface. With the tendency to form micelles, determination of the Kow for a surfactant will also be strongly dependent on its concentration (Müller et al. 1999b). Instead, the C logP based on the structure of the chemical can be determined using a modified version of the fragment method of Leo and Hansch (Roberts 1991; Roberts and Marshall 1995). In this method, using
linear C12EO1 with a logP of 5.23 as the parent compound, an increment of 0.54 is added for each additional carbon unit, and an increment of –0.10 is used for each additional EO group. To adjust for branching in the carbon chain, the term [1.44 log (Cs + 1)] is subtracted, where Cs is the carbon number of the shorter chain (Roberts and Marshall 1995).
Next, for each specific homologue in question, the appropriate QSAR equation found in Table 4 was used to predict the toxicity for each of the available toxicity endpoints listed in Table 2 (for 17 species = three fish species, eight invertebrate species and six plant species) given the toxicity of the original AE tested. A species sensitivity distribution was then constructed for each specific homologue. The HC5 from the species sensitivity distribution was determined as the FWQG for each specific AE homologue which appears in each cell of Table 3.
Table 4. Quantitative Structure Activity Relationships (QSARs) used to convert toxicity data shown in Table 1 for default homologue to homologues shown in Table 3
|QSAR Equation||Units||Type of Data to Which the QSAR is Applied||Reference|
|log (72h EC20 algae) = -0.378 × log Kow –4.072||mmol·L-1||Chronic Plant/Algae Toxicity||Wind and Belanger 2006|
|log (21d EC20 Daphnia magna, reproduction) = -0.532 × log Kow – 3.025||mol·L-1||Chronic Invertebrate Toxicity||Boeije et al. 2006|
|log (30d EC20 Pimephales promelas) = -0.307 × log Kow – 3.92||mol·L-1||Chronic Fish Toxicity||Boeije et al. 2006|
|log (mesocosm NOEC) = -0.74 x logKow –2.78||mol·L-1||Mesocosm||Boeije et al. 2006|
|log (48h EC50 Daphnia magna) = -0.38 × alkyl-length + 0.1 × EO-length – 1.77||mol·L-1||Acute Invertebrate Toxicity||Wong et al. 1997|
|log (96h LC50 Pimephales promelas) = -0.34 × alkyl-length + 0.05 × EO-length -1.65||mol·L-1||Acute Fish Toxicity||Wong et al. 1997|
Considerations in Guideline Implementation
The default FWQG for the protection of freshwater life from adverse effects of AEs is 0.07 mg/L normalized to the average homologue distribution of C13.7EO5. Together with the other values shown in Figure 1, these guidelines can be used when the average homologue distribution of AEs at a particular site are unknown. A simple measurement of total AE in the ambient waters can be compared with the guideline values to determine whether there is an exceedence. This is the most simple and cost effective approach, as chemical analyses for individual homologues can be costly, and very few analytical laboratories are currently set up to do them. Although analyses of individual homologues are not required to use the default guidelines, the analytical method used should be capable of detection at low ppb concentrations, as well as detection of free fatty alcohols and AEs with short EO chains.
The default guidelines are based on an environmental distribution expressed for convenience as a C13.7EO5 AE, which is likely characteristic of many sites where the main source of AEs is through municipal wastewater effluents. Where AEs in the ambient water are originating from other sources, such as industrial outflows, the average homologue distribution may be considerably different. If an initial measurement of total AE at the site has been found to exceed the default guidelines, then more in-depth evaluation through analysis of specific homologues could be considered. Use of the homologue-specific guidelines will give a better indication of whether AEs present at a particular site are in fact likely to cause adverse effects or not. In these cases, it would be advisable to use the homologue-specific guideline values presented in Table 3.
For determining the concentrations of individual AE homologues in ambient water at a particular site, the current recommended analytical approach is the pyridinium LC/MS method (Dunphy et al. 2001). Concentrations should be determined, at a minimum, for homologues with alkyl chains between 12 and 18 carbons long (excluding C17 AEs), and ethoxylate chains ranging from 0 to 18 EO units long. The measured concentrations of the homologues can then be compared with the homologue -specific guideline values in Table 3 using a Toxic Unit approach (see Boeije et al. 2006 for additional background on the construction of Toxic Unit approaches for AEs). For each individual homologue, the measured concentration is divided by the corresponding homologue -specific guideline listed in Table 3 to obtain the toxic unit for that particular homologue. The toxic units for all homologues present in the sample are then summed. If the total of all toxic units is less than or equal to one, then the AE guideline has not been exceeded. If the total of all toxic units is greater than one, this indicates an exceedence of the guideline.
For alcohol alkoxylates, where the hydrophilic chain of the surfactant contains a mix of ethoxylate, propoxylate, and/or butoxylate groups, the FWQGs for AEs may also be applied. This recommendation is based on data (see Roberts et al. 2007) that suggest that alcohol alkoxylates have similar aquatic toxicity to AEs of the same parent alcohol and with similar numbers of ethoxylate groups.
Aveyard, R. 1984. Adsorption at the air/liquid, liquid/liquid, and solid/liquid interfaces. Pages 153-173 in T. F. Tadros, ed. Surfactants. Toronto: Academic Press, Inc.
Boeije, G.M., M.L. Cano, S.J. Marshall, S.E. Belanger, R. Van Compernolle, P.B. Dorn, H. Gümbel, R. Toy, and T. Wind. 2006. Ecotoxicity QSARs for alcohol ethoxylates based on the mixture toxicity concept. Ecotox. Environ. Safe. 64:75-84.
Birch, R.R. 1984. Biodegradation of nonionic surfactants. J. Am. Oil. Chem. Soc. 61: 340-343.
Bishop, W.E. and A.W. Maki. 1980. A critical comparison of two bioconcentration test methods. Pages 61-77 in J. G. Eaton, P.R. Parrish, and A.C. Hendricks, eds. Aquatic Toxicology, ASTM STP 707. American Society for Testing and Materials.
Bishop, W.E. and R.L. Perry. 1981. Development and evaluation of a flow-through growth inhibition test with duckweed (Lemna minor). Pages 421-435 in D.R. Branson and K.L. Dickson, eds. Aquatic Toxicology and Hazard Assessment, Proceedings of the Fourth Annual Symposium on Aquatic Toxicology. Philadelphia, PA: American Society for Testing and Materials.
Camford Information Services, Inc. 1997. CPI Product Profiles: Ethoxylated alcohols (linear alcohol ethoxylates, etc). Toronto. 2 pp.
Campbell, P.M. 2002. Alternatives to nonylphenol ethoxylates: Biodegradation, and technical-economic aspects. Report prepared for Environment Canada. 83 pp. Cardellini, P. and L. Ometto. 2001. Teratogenic and toxic effects of alcohol ethoxylate and alcohol ethoxy sulfate surfactant on Xenopus laevis embryos and tadpoles. Ecotox. Environ. Safe. 48: 170-177.
CCME (Canadian Council of Ministers of the Environment). 2007. A Protocol for the Derivation of Water Quality Guidelines for the Protection of Aquatic Life. In: Canadian Environmental Quality Guidelines, 1999, Canadian Council of Ministers of the Environment, Winnipeg.
Dorn, P.B., J.H. Rodgers, S.T. Dubey, W.B. Gillespie, Jr., and R.E. Lizotte, Jr. 1997a. An assessment of the ecological effects of a C9-11 linear alcohol ethoxylate surfactant in stream mesocosm experiments. Ecotoxicology 6: 275-292.
Dorn, P.B., J.H. Rodgers, Jr., W.B. Gillespie, Jr., R.E. Lizotte, Jr., and A.W. Dunn. 1997b. The effects of a C12-13 linear alcohol ethoxylate surfactant on periphyton, macrophytes, invertebrates and fish in stream mesocosms. Environ. Toxicol. Chem. 16: 1634-1645.
Dorn, P.B., J.P. Salanitro, S.H. Evans, and L. Kravetz. 1993. Assessing the aquatic hazard of some branched and linear nonionic surfactants by biodegradation and toxicity. Environ. Toxicol. Chem. 12: 1751-1762.
Dunphy, J.C., D.G. Pessler, and S.W. Morrall. 2001. Derivatization LC/MS for the simultaneous determination of fatty alcohol and alcohol ethoxylate surfactants in water and wastewater samples. Environ. Sci. Technol. 35: 1223-1230.
Eadsforth, C.V., A.J. Sherren, M.A. Selby, R. Toy, W.S. Eckhoff, D.C. McAvoy,a nd E. Matthijs. 2006. Monitoring of environmental fingerprints of alcohol ethoxylates in Europe and Canada. Ecotox. Environ. Safe. 64(1): 14-29.
Garcia, M.T., I. Ribosa, S. Leal, and W. Hreczuch. 1996. Effect of homologue distribution on the toxicity of alcohol ethoxylates. J. Am. Oil. Chem. Soc. 73: 903-906.
Ghirardini, A.V., A.A. Novelli, B. Likar, G. Pojana, P.F. Ghetti, and A. Marcomini. 2001. Sperm cell toxicity test using sea urchin Paracentrotus lividus Lamarck (Echinodermata: Echinoidea): Sensitivity and discriminatory ability toward anionic and nonionic surfactants. Environ. Toxicol. Chem. 20: 644-651. Government of Canada. 1999. Canadian Environmental Protection Act, 1999. S.C., 1999, c. 33, Canada Gazette, Part III: Vol. 22, No. 3.
Government of Canada. 2000. Canadian Environmental Protection Act, 1999: Persistence and Bioaccumulation Regulations, P.C. 2000-348, 29 March, 2000, SOR/2000-107. Canada Gazette, Part II: Vol. 134, No. 7.
Hansch, C., J.E. Quinlan, and G.L. Lawrence. 1968. Linear free-energy relationship between partition coefficients and the aqueous solubility of organic liquids. J. Organic Chem. 33: 347-50
Hennes-Morgan, E.C. and N.T. De Oude. 1994. Detergents. Pages 130-153 in P. Calow, ed. Handbook of Ecotoxicology. Boston: Blackwell Scientific Publications, Inc.
Holt, M.S., G.C. Mitchell, and R.J. Watkinson. 1992. The environmental chemistry, fate and effects of nonionic surfactants. Pages 89-144 in N.T. De Oude, ed. Anthropogenic compounds: Detergents. Berlin / Heidelberg, Germany: Springer-Verlag.
Kaluza, U. and K. Taeger. 1996. Einfluß der chemischen struktur auf ökotoxikologische eigenschaften von alkanol-ethoxylaten. Tenside Surf. Det. 33: 46-51.
Kiewiet, A.T., J.R. Parsons, and H.A.J. Govers. 1997. Prediction of the fate of alcohol ethoxylates in sewage treatment plants. Chemosphere 34: 1795-1801.
Kurata, N., K. Koshida, and T. Fujii. 1977. Biodegradation of surfactants in river water and their toxicity to fish. Yukagaku 26: 115-118.
Lewis, M.A. 1990. Chronic toxicities of surfactants and detergent builders to algae: A review and risk assessment. Ecotox. Environ. Safe. 20: 123-140.
Lewis, M.A. and B.G. Hamm. 1986. Environmental modification of the photosynthetic response of lake plankton to surfactants and significance to a laboratory-field comparison. Wat. Res. 20: 1575-1582.
Lizotte, R.E., Jr., D.C.L. Wong, P.B. Dorn, and J.H. Rodgers, Jr. 1999. Effects of a homologous series of linear alcohol ethoxylate surfactants on fathead minnow early life stages. Arch. Environ. Contam. Toxicol. 37: 536-541.
Macek, K. and S.F. Krzeminski. 1975. Susceptibility of bluegill sunfish (Lepomis macrochirus) to nonionic surfactants. Bull. Environ. Contam. Toxicol. 13: 377-384.
Madsen, T., H. Buchardt Boyd, D. Nylén, A. Rathmann Pedersen, G.I. Petersen, and F. Simonsen. 2001. Environmental and Health Assessment of Substances in Household Detergents and Cosmetic Detergent Products. Environmental Project No. 615. Danish EPA. 240 pp.
Maki, A.W. and W.E. Bishop. 1979. Acute Toxicity of Surfactants to Daphnia magna and Daphnia pulex. Arch. Environm. Contam. Toxicol. 8: 599-612.
Marcomini, A., G. Pojana, C. Carrer, L. Cavalli, and G. Cassani. 2000. Aerobic biodegradation of monobranched aliphatic alcohol polyethoxylates. Environ. Toxicol. Chem. 19: 555-560.
Matthijs, E., G. Debaere, N. Itrich, P. Masscheleyn, A. Rottiers, M. Stalmans, and T. Federle. 1995. The fate of detergent surfactants in sewer systems. Wat. Sci. Tech. 31: 321-328.
McAvoy, D.C., S.D. Dyer, N.J. Fendinger, W.S. Eckhoff, D.L. Lawrence, and W.M. Begley. 1998. Removal of alcohol ethoxylates, alkyl ethoxylate sulfates, and linear alkylbenzene sulfonates in wastewater treatment. Environ. Toxicol. Chem. 17: 1705-1711.
McAvoy, D.C. and K.M. Kerr. 2001. Association of alcohol ethoxylates with a dissolved humic substance. Pages 177-186 in Humic Substances and Chemical Contaminants. Madison, WI: Soil Science Society of America.
Modler, R.F., R. Gubler, and A. Kishi. 2002. Surfactants, household detergents and their raw materials. Chemical Economics Handbook. CEH Marketing Report. 89 pp.
Modler, R.F., R. Gubler, and Y. Inoguchi. 2004. Detergent alcohols. In: Chemical Economics Handbook. SRI International Consulting, Menlo Park, pp. 609.500A – 609.500S. (as cited in Federle and Itrich 2006)
Moore, S.B., R.A. Diehl, J.M. Barnhardt, and G.B. Avery. 1987. Aquatic toxicities of textile surfactants. Text. Chem. Color 19: 29-32.
Mudge, S.M., S.E. Belanger, and A.M. Nielsen. 2008. Fatty Alcohols and Natural Occurrence in the Environment. Royal Society of Chemistry, Cambridge, UK. 159 pp.
Mulla, M.S., H.A. Darwazeh, and L.L. Luna. 1983. Monolayer films as mosquito control agents and their effects on nontarget organism. Mosq. News 43: 489-495.
Müller, M.T., A.J.B. Zehnder, and B.I. Escher. 1999a. Membrane toxicity of linear alcohol ethoxylates. Environ. Toxicol. Chem. 18: 2767-2774.
Müller, M.T., J.B. Zehnder, and B.I. Escher. 1999b. Liposome-water and octanol-water partitioning of alcohol ethoxylates. Environ. Toxicol. Chem. 18: 2191-2198.
Newsome, C. S., D. Howes, S. J. Marshall, and R. A. van Egmond. 1995. Fate of some anionic and alcohol ethoxylate surfactants in Crassius auratus. Tenside Surf. Det. 32: 498-503.
Nielsen, A.M., A.J. DeCarvalho, D.C. McAvoy, L. Kravetz, and M.L. Cano. 2002. Investigation of an onsite wastewater treatment system in sandy soil: site characterization adn fate of anionic and nonionic surfactants. Environ. Toxicol. Chem. 21: 2606-2616.
Nyberg, H. 1988. Growth of Selenastrum capricornutum in the presence of synthetic surfactants. Wat. Res. 22: 217-223.
P&G - Shell. 2003. Preliminary summary evaluation of alcohol ethoxylates in Canadian effluents. Prepared for Environment Canada by The Procter & Gamble Company and Shell Global Solutions. 13 pp.
Raney, K.H. 2000. Impact of phase behavior on aquatic toxicity testing of alcohol ethoxylates. Colloid. Surface. A 167: 151-164.
Roberts, D.W. 1991. QSAR issues in aquatic toxicity of surfactants. Sci. Total Environ. 109/110: 557-568.
Roberts DW, Marshall SJ. 1995. Application of hydrophobicity parameters to prediction of the acute aquatic toxicity of commercial surfactant mixtures. SAR and QSAR in Environmental Research 4:167–176.
Roberts, J.F., S.J. Marshall, and D.W. Roberts. 2007. Aquatic toxicity of ethoxylated and propoxylated alcohols to Daphnia magna. Environ. Toxicol. Chem. 26: 68-72.
Sherren, A.J., M.A. Selby, and C.V. Eadsforth. 2003. Analysis of effluents for alcohol ethoxylates from a Canadian monitoring study. Analytical Technology Business Group, Shell Global Solutions (UK). Chester, England. 24 pp.
Stephenson, R.J., C.V. Eadsforth, A.J. Sherren, J.D. Williams, M.A. Selby, and R.N. Bumpus. 2004. NEODOL 25-7 and GENAPOL T110: an assessment of the fate of a 2:1 alcohol ethoxylate mixture during simulated aerobic sewage treatment. RTS Report OG.03.49009. Shell Global Solutions International B.V., The Hague, The Netherlands. 74 pp.
Swisher, R.D. 1987. Surfactant Biodegradation. New York: Marcel Dekker, Inc.
Talmage, S. 1994. Environmental and Human Safety of Major Surfactants: Alcohol Ethoxylates and Alkylphenol Ethoxylates. Lewis Publishers, Boca Raton, FL. 374 pp.
Tolls, J., M. Haller, E. Labee, M. Verweij, and D.T.H.M. Sijm. 2000. Experimental determination of bioconcentration of the nonionic surfactant alcohol ethoxylate. Environ. Toxicol. Chem. 19: 646-653.
Tolls, J., P. Kloepper-Sams, and D.T.H.M. Sijm. 1994. Surfactant bioconcentration -- a critical review. Chemosphere 29: 693-717.
Tovell, P.W.A., C. Newsome, and D. Howes. 1975. Effect of water hardness on the toxicity of a nonionic detergent to fish. Wat. Res. 9: 31-36.
van Compernolle, R., D.C. McAvoy, A. Sherren, T. Wind, M.L. Cano, S.E. Belanger, P.B. Dorn, and K.M. Kerr. 2006. Predicting the sorption of fatty alcohols and alcohol ethoxylates to effluent and receiving water solids. Ecotoxicol. Environ. Safety 64: 61-74.
Vashon, R.D. and B.S. Schwab. 1982. Mineralization of linear alcohol ethoxylates and linear alcohol ethoxy sulfates at trace concentrations in estuarine water. Environ. Sci. Technol. 16: 433-436.
Wakabayashi, M., M. Kikuchi, A. Sato, and T. Yoshida. 1987. Bioconcentration of alcohol ethoxylates in Carp (Cyprinus carpio ). Ecotox. Environ. Safe. 13: 148-163.
Wind, T. and S.E. Belanger 2006. Acute and chronic toxicity of alcohol ethoxylates to the green alga, Desmodesmus (=Scenedesmus) subspicatus ,and the subsequent development of structure activity relationships. Bull. Environ. Contam. Toxicol. (2006) 76:218–225.
Wong, D.C.L., P.B. Dorn, and E.Y. Chai. 1997. Acute toxicity and structure-activity relationships of nine alcohol ethoxylate surfactants to fathead minnow and daphnia magna. Environ. Toxicol. Chem. 16: 1971-1976.
Yamane, A.N., M. Okada, and R. Sudo. 1984. The growth inhibition of planktonic algae due to surfactants used in washing agents. Wat. Res. 18: 1101-1105.
List of Acronyms
- alcohol ethoxylates
- Chemical Abstracts Service
- Canadian Council of Ministers of the Environment
- Chemical Management Plan
- dissolved organic matter
- effects concentration
- ethylene oxide
- Federal Water Quality Guidelines
- lethal concentration
- maximum acceptable toxicant concentration
- quantitative structure activity relationship
- species sensitivity distribution
- wastewater treatment plant
- Date modified: