Cleanup: Volume VI: The Remainders Collection


This compilation covers various homologues and test matrixes, including C 12 EO 10 in activated sludge Kiewiet et al. These sorption coefficients can be used in an aquatic risk assessment to account for the bioavailability of each homologue, if there are enough data to estimate K d values for each of the homologues of AE. Since it is impractical to measure all of these K d values, a quantitative correlation of carbon chain length C and ethoxylate number EO based on the existing data was developed by van Compernolle et al.

This relationship illustrates that AE sorption is mostly controlled by the alkyl chain length, where an increase in alkyl chain length causes an increase in sorption. An increase in the ethoxylate number has only a slight negative effect on AE sorption. This slight effect may be due in part to the fact that the test matrices used in these studies have high organic carbon content which results in more sites for hydrophobic interaction.

Thus, this model may not be appropriate for other types of matrices where the organic matter content is lower and the clay content is higher e. Because of this limitation, the authors suggest that this model should only be used when the fraction of organic carbon f oc is greater than 0. The resulting K d predictions for each homologue were used to estimate the bioavailability adjustment in the exposure concentrations as part of the aquatic risk assessment of AE Belanger et al.

Alkyl sulfates also known as alcohol sulfates AS are produced by sulfation of detergent range primary alcohols Figure 1 using sulfur trioxide or chlorosulfonic acid followed by neutralization with a base. The most common neutralizing agent used is a sodium salt, less commonly an ammonium salt and very minor volumes are neutralized with alkanolamines, usually TEA resulting in the sodium, ammonium, or TEA salts, respectively. Commercial grades of linear-type primary AS are typically in the C 12 —C 18 range. AS are used in household cleaning products such as laundry detergents, hand dishwashing liquids, and various hard-surface cleaners, personal care products, institutional cleaners, and industrial cleaning processes, and as industrial process aids in emulsion polymerisation and as additives during plastics and paint production HERA, This represents a sharp decline from the peak of This decline was largely due to the declining use of powder laundry detergents, which contain AS, as consumers switched to liquid laundry detergents.

Demand for AS in powder laundry detergent use has continued to decline and reached Overall, consumption of detergent alcohols to make AS is expected to decline at a rate of 3. When used in personal care products, this surfactant is mostly based on C 12 —C 14 alcohols. The largest remaining applications of AS are in emulsion polymerization and as emulsifiers for agricultural herbicides. The number of carbon units in the AS affects the surfactants physical and chemical as well as its partitioning and fate properties in the environment.

Table 4 summarizes the core physical and chemical properties of different AS chain lengths assuming these are sodium salts. Note that the water solubility decreases dramatically with increasing carbon chain length. The relatively high water solubility combined with its surfactant properties explains why AS 12 is the most widely used AS in detergents. Physical and chemical properties of AS surfactants of various carbon chain lengths assuming sodium salt. Numerous screening level tests have been conducted to evaluate the aerobic biodegradation of AS.

Rapid biodegradation of C 12 AS was also observed in river water Guckert et al. Kikuchi and Knaggs et al. A major factor that affects the rate of biodegradation is the linearity of the alkyl chain although slight branching of the alkyl chain does not hinder the biodegradation of AS Battersby et al. In contrast, some highly branched AS homologues have been observed to degrade at a much slower rate SDA, c. Temperature has little effect on the rate of biodegradation in activated sludge Gilbert and Pettigrew, and river water Lee et al.

The anaerobic biodegradation of AS has also been investigated. Branching of the alkyl chain reduces the extent of ultimate anaerobic biodegradation Rehman et al. More definitive biodegradation tests using radiolabeled 14 C linear AS at realistic concentrations have been conducted with anaerobic digester sludge. Sorption distribution coefficients K d for several AS homologues C 8—14 have been reported for two river sediments Marchesi et al.

All of the AS homologues exhibited fast adsorption to the river sediments less than 20 min. An extensive oxidative treatment of the sediments greatly reduced the sorption capacity for AS, suggesting a hydrophobic mechanism of interaction. Measured K d values increase as the chain length of the AS increases. AES are essentially ethoxylated AS where the carbon chain length, ranges from 12 to 18 and the number of ethoxylate groups, ranges from 0 to 8 Table 1. The alkyl chain is ethoxylated to a predetermined average number of EO groups and sulfated to provide a product with the desired properties Biermann et al.

AES are produced by sulfation of the ethoxylates of primary alcohols Figure 1 , using sulfur trioxide or chlorosulfonic acid followed by immediate neutralization with base to produce typically a sodium salt, less commonly an ammonium salt SRI Consulting, a. Minor volumes are neutralized with alkanolamines, usually TEA. The commercially produced AES can contain a mixture of as many as 36 homologues with the actual composition reflecting the aliphatic alcohol feedstock selection and the average degree of desired ethoxylation. Most commercial AES are produced as low or high aqueous active solutions, e.

AES are a widely used class of anionic surfactants. These are used in household cleaning products such as laundry detergents, hand dishwashing liquids, and various hard-surface cleaners, personal care products, institutional cleaners, and industrial cleaning processes, and as industrial process aids in emulsion polymerization and as additives during plastics and paint production HERA, AES use in household cleansers is expected to grow as a result of the growth in use of germicidal disinfectants, which often contain AES.

AES is less irritating to the skin and eyes than many other surfactants, so the use of AES in personal cleansing products is also expected to increase since there has been a general trend toward milder personal care products SRI Consulting, a. Since the mids, there has been consistent growth in germicidal disinfectants used to clean household kitchen counters; these products often contain AES.

Along with institutional and commercial cleaning, the largest applications are emulsion polymerization and emulsifiers for agricultural herbicides. AES are anionic surfactants and share many of the same trends in physical and chemical properties with other anionic surfactants, especially AS Table 5. The ethoxylation process increases the size and weight of the molecule compared to AS which slightly increases their water solubility compared to an AS of the same carbon chain length.

Several screening level tests have been conducted to evaluate the aerobic biodegradation of AES. As a class of compounds, linear AES used in detergent products alkyl chain length C 12—16 and ethoxylate chain length EO 1—4 undergo rapid primary and ultimate biodegradation Kravetz et al. Neither the length of the alkyl chain i.

A major factor that affects the rate of biodegradation is the linearity of the alkyl chain.

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Some highly branched AES homologues have been observed to degrade at a much slower rate in a river water die-away test Yoshimura and Masuda, Little published information is available on the anaerobic biodegradation of AES. However, based on the chemical structure of AES and the rapid anaerobic biodegradability of the structurally related AE and AS, the biodegradability of AES in anaerobic environments is expected Steber and Berger, An anaerobic screening biodegradability test by Steber supports this conclusion.

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Low anaerobic biodegradation potential for AES has been reported in some cases Madsen and Rasmussen, ; Fraunhofer, The low gas production in these tests can be attributed to the very high test substance to biomass ratio used. Gilbert and Pettigrew reported that AES to biomass ratios of 0. Little published information is available on the sorption of AES to environmental surfaces. They found the amount of AES sorbed was strongly correlated with the organic carbon content of the sediments. LAS is an anionic surfactant containing a hydrophobic region alkyl carbons and the phenyl group and a hydrophilic group the sulfonate group as shown in Table 1.

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The sulfonate group is situated para to the alkyl group and the alkyl group generally contains 10—14 carbons. The attachment of the phenyl group to the alkyl carbons occurs at any interior alkyl carbon, and the phenyl position is referred to as the carbon number i. The average chain length of commercial LAS is approximately Most commercial LAS products are mixtures of isomers and homologues. In the Friedel—Crafts reaction, n -paraffins are dehydrogenated to form the n -olefin that is combined with benzene, typically in the presence of an AlCl 3 or HF catalyst to form the alkyl benzene de Almeida et al.

Use of the HF catalyst gives an even distribution of phenyl position along the n -paraffin chain between C-2 and C-6 positions i. To minimize the formation of impurities, manufacturers preferentially used the HF catalyst in the s and early s. In this process, HF and AlCl 3 catalysts are replaced with various solid acid catalyst-based systems e.

The Detal process is cost-effective, generates LAS enriched in the 2-phenyl positional isomer, with greater linearity of the alkyl chain, and lower levels of impurities in the LAB, such as dialkyl tetralin sulfonates than the HF-catalyzed process Kocal et al. Currently, the most common sulfonation process uses a falling film reactor with and SO 3 gas. Sulfonation of LAB generates alkylbenzene sulfonic acid, which is then neutralized with a base to give the final LAS surfactant salt. Sodium-neutralized LAS are most common but other materials can be used to give the resulting LAS salts other beneficial properties.

LAS is the world's largest-volume synthetic surfactant with over 4 million metric tons consumed worldwide in SRI Consulting, b. From the late s until the early s, LAS was the largest volume surfactant manufactured and consumed in household detergents in North America.

At its peak production in the early s, North American production and use was approximately , metric tons. This decline was due to increases in LAS prices driven by higher raw material costs, lower surfactant levels in products as a result of increased enzyme use, and replacement of LAS by AES. LAS are widely used in a variety of detergent formulations including laundry powers and liquids, dishwashing liquids, car washes, and hard-surface cleaners SRI Consulting, b.

Due to its strength as a cleaning agent, LAS is not often used in personal care products. Industrial and institutional detergents and cleaners also rely heavily on LAS, and it is also used as an emulsifier e. Very small volumes are also used in personal care applications. Thus, almost , metric tons of LAS were consumed in North American household detergents in with a peak consumption of over , metric tons in the early s. Since LAS is a mixture of homologues and isomers, a range of values for any one property is expected Table 6. If the phenyl position is kept constant, as the chain length increases, then the hydrophobicity will increase resulting in an increase in K ow and a decrease in solubility.

The effect of chain length on a physical parameter can be substantial. Physical chemical data of LAS by calculated methods based on the pure homologue, 2-phenyl isomer. The octanol—water partition coefficient, log K ow , cannot be experimentally measured for surfactants because of their surface-active properties, but can be approximated using various estimation methods such as Roberts This value was used in the aquatic risk assessment carried out in the Netherlands Feijtel and van de Plassche, Numerous screening level tests have been conducted to evaluate the aerobic biodegradation of LAS.

As a class of compounds, LAS undergoes rapid primary and ultimate biodegradation, and is classified as readily biodegradable Swisher, ; European Union Commission, While the day window is no longer necessary for assessing the ready biodegradability of surfactants CSTEE, , several studies have reported that LAS meets the day window. Higher tier tests have also shown that the biodegradation intermediates sulfo phenyl carboxylates SPC are not persistent Gerike and Jasiak, ; Cavalli et al. In a more definitive study, Itrich and Federle used radiolabeled 14 C LAS to determine a first-order primary biodegradation rate of 0.

Field studies have demonstrated in-stream half-life losses for LAS in the range of 1—3 hr, though some of this loss could be due to sorption and settling to the river bed Takada et al. In a seawater biodegradation test, Vives-Rego et al. The anaerobic biodegradation of LAS has also been investigated. Several laboratory screening tests, which determine ultimate biodegradation by measuring gas production CH 4 and CO 2 over a two month incubation period, did not show significant anaerobic biodegradation of LAS Steber, ; Federle and Schwab, ; Gejlsbjerg et al.

Environmental Safety of the Use of Major Surfactant Classes in North America

Based on these studies, it is generally recognized that LAS is not biotransformed in anaerobic environments, though under oxygen-limited field conditions biodegradation of LAS can be initiated and then continue in anaerobic environments Larson et al. Sulfate-reducing bacteria, firmicutes, and clostridia were identified as possible candidates for causing the degradation. K d sediment values were higher than K d soil ones, as a consequence of the higher organic content in sediment than in soil Marchesi et al.

The LAS sorption distribution coefficients K d can vary greatly due to the structural variability of LAS mixture of homologues having alkyl chain lengths ranging from C 10 —C 14 with isomers having phenyl positions ranging from 2 to 7 , aqueous solubility of the homologues ranging from 0.

In another study with river sediments, Marchesi et al. Differences in the K d values between these studies could be due to the distribution of homologues and phenyl positions or the characteristics of the sediment. In a study investigating the sorption of LAS to river suspended solids, Belanger et al. An investigation on the association of LAS with dissolved humic acids by Traina et al. To understand the fate and exposure of surfactants and LCOHs present in various consumer products, one needs to understand the typical pathways that these chemicals take to enter the environment following use in the household and the fate processes that affect their concentrations during transit.

Figure 3 illustrates the typical pathways in North America for these types of ingredients to reach the environment after household use. While exposure can be determined by direct measurement in environmental compartments of interest, such measurements represent the exposure only at a specific time and site situation because chemical usage patterns, wastewater flow rates, wastewater removal efficiencies, and surface water flow rates can vary with time and from location to location.

An alternate approach is to model exposure based upon first principles for specific or generic scenarios Cowan et al. However, this approach is limited by the availability of data to parameterize the model. The use of model predictions to estimate the exposure in the environment along with field measurements provides a high level of confidence that real-world exposures are being addressed in the risk assessment. Therefore in this paper, we will include both the predicted exposure based on mathematic models Section IV and field measurements to dimension the exposure of these chemicals in the risk assessments Section VI.

Only a brief overview of the exposure calculations and models will be provided here. The reader is referred to other papers and books for more details Cowan et al. Because these chemicals are used predominantly in down-the-drain consumer and industrial products such as laundry detergents, dishwashing detergents, and personal cleansing, the first step in the pathway of these chemicals to the environment is their release from the household into the wastewater conveyance systems Fig 3.

The equation used to calculate their concentration in household wastewater, C ww is:. The AMT is typically estimated from the tonnage of the chemical sold over a year in the country or region of interest such as the data summarized in Section II divided by the number of people in the population, P , and the number of days in a year i. Default values chosen for P are 3. Census Bureau, , combined with 3. Therefore, the next step in the pathway to the surface water environment is transport in sewer conveyance systems.

Although sewers were once thought to be just conveyance systems, several studies have demonstrated that these actually serve as bioreactors Matthijs et al. The approach for incorporating sewer loss into the exposure assessment is to treat sewer conveyance as a completely mixed reactor with a first-order loss rate in wastewater under conditions representative of a sewer. Typically sewers have low but not fully anoxic, dissolved oxygen levels approximately 0.

The average concentration in the sewer would be the influent concentration to the WWTP but if degradation has occurred, the influent concentration will be less than the concentration in the wastewater from the home estimated in Equation 1. The average concentration in the sewer, C sew , can be calculated using the following equation:.

If information on HRT does not exist, then loss in the sewer can conservatively be assumed to be zero. Alternatively, as is done here see Section IV. B , the loss rate in the sewer can be estimated from the measured WWTP influent concentration compared to the estimated household wastewater concentration.

Lower concentrations in the influent than those estimated from household use could also occur if the household wastewater is diluted with non-household wastewater discharged to the same sewers; therefore, to estimate the loss rate in sewers the presence of non-household discharges must be minimal. As will be discussed in Section IV. B, the measured influent concentrations in Sanderson et al. A Dyer, personal communication. For example, the ratio of measured to estimated influent values of AES was 0.

The main site for removal of the chemicals before entering the environment is in WWTPs Fig 3 by sorption on sewage sludge and loss via biodegradation. WWTP operational parameters e. For activated sludge plants, the effects of these operational parameters can be accounted for by wastewater simulation models Struijs et al. The prediction of effluent concentration, C effluent , and sludge concentrations, C sludge , will depend on the treatment plant operational parameters.

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The bulk of WWTP effluents are released into surface waters. At the local scale, surface water concentrations at the point of the effluent discharge, C surface water , can be calculated by:. Although single default dilution factors are commonly used U. A default value for DF of 1, which represents no dilution of the WWTP effluent can be used to provide a conservative estimate of the surface water concentration. Once the WWTP effluent is diluted into surface waters, the processes of sedimentation and biodegradation will act to further reduce the C surface water.

The fate of the specific chemical will be dependent on the residence time of that chemical in the surface water, its sorption, and degradation properties, presence and type of suspended solids, sedimentation of solids, and presence of an active microbial community. These factors may be considered in the calculation of downstream surface water concentrations, C downstream , by:. Clearly, one of the chief determinants here is the duration of the travel time downstream from the point of discharge. Any of the chemical that is attached to particles can become incorporated into sediment due to settling of these particles.

A chemical's concentration in the sediment depends on the sorption constant to the suspended and sediment solids and the concentration of these solids in the water column and the sediment. The two possible calculation methods are provided below to estimate the sediment concentration. The first is explained in detail in Cowan et al. The first step is to calculate the dry weight concentration of the chemical on the suspended particles in the water column, C ss , using the following equation:.

Alternatively, the sediment concentration can be estimated from the surface water concentration, the organic carbon partition coefficient, K oc , and the organic carbon content of the sediment i. This approach assumes that the sediment is in equilibrium with the overlying water which is not unreasonable for surface sediment. To conduct the prospective risk assessment of the five chemical classes, the concentrations of these chemicals in surface waters and sediments are estimated using the approach described in Section III and the specific physical, chemical, and degradation data for each surfactant in Section II.

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Section II of the paper presents these environmental exposure estimates. B Dyer, personal communication. This ratio of measured to estimated influent values from Sanderson et al. As described previously, to predict the concentrations in surface waters, the next step is to determine the removal of the surfactants in wastewater treatment systems. Data are available from monitoring studies at a wide range of wastewater treatment systems for these surfactants. Alkyl chain length and isotope signatures of carbon and hydrogen have been used to assess the removal of LCOHs as well as to distinguish among the sources of alcohols i.

Recent studies by Mudge and Mudge et al. In effluent, the remaining alcohols have a signature that is unlike influent, as a result of mixed-liquor in situ bacterial synthesis of alcohols. Therefore, LCOHs measured in surface water and sediments are from natural in situ synthesis or terrestrial runoff sources not detergents. A summary of U. Monitoring data and removal values for long-chain alcohols in U. Also, the removal values are based on measured total concentrations which include LCOH.

For PT treatment, the removal was estimated based on the first stage of three activated sludge and five tricking filter plants. The removal of AE during PT treatment ranged from Monitoring data and removal values for AE in U. The average TFT removal values ranged from For these three treatment types, the average removals ranged from In the United States, one study by Fendinger et al. An average removal for the ODT treatment plant was One study McAvoy et al. Also, the removal values are based on total concentrations, which include AS.

The average influent, effluent, and removal values for the AST treatment plants were 0. For TFT treatment, the average influent, effluent, and removal values were 0. A large number of monitoring studies have been conducted in the United States to determine the removal of LAS in various types of wastewater treatment Table Monitoring data and removal values for LAS in U. The overall average removal from these three studies was The average removal value was This information, along with effluent dilution by the receiving stream which is calculated from the WWTP flow and the receiving water flow, is used to predict receiving water concentrations under either mean or low-flow 7Q10 conditions for all U.

The 7Q10 values represent the lowest 7-day average flow in a year that occurs during 7 consecutive days on average once every 10 years. The input parameters used to predict receiving water exposure concentrations for each of the surfactants are provided in Table The volume of the chemicals and the population numbers used to estimate the per capita use are based on As indicated in Table 11 , when WWTP removal values were not available for a particular treatment plant type e.

An in-stream loss rate of 0. Because of this, predicted exposure concentrations are used when assessing ecological risk in the prospective risk assessment Section VI. The objective of this section is to describe how the predicted no-effect concentration PNEC for each of the surfactants is derived. The mechanism of toxicity for surfactants is accepted to be nonpolar narcosis in which the surfactant's presence in the cell membrane is believed to interfere with membrane-dependent life processes such as energy metabolism and transport of nutrients and oxygen across the membrane.

For example, the toxicity of several anionic and nonionic surfactants has been observed to be highly correlated with properties like standard free energies and interfacial activity Rosen et al. Both the nature of the hydrophobe the alkyl chain and the nature of the hydrophilic head group contribute to defining the magnitude of these properties for a specific surfactant homologue. The hydrophobe determines the ease with which the surfactant will insert itself into the membrane bilayer and the amount of disturbance due to hydrophobic interactions caused once it is in the membrane.

This explains the observed pattern, typical for the ecotoxicity of surfactants in general, of increasing toxicity with increasing alkyl chain length—until a point is reached where water solubility becomes the limiting factor. For several of the very long-chain length surfactants, solubility is greatly reduced and thus the toxicity decreases, especially for hydrocarbon chain lengths of 15 and above.

Bernhard and Dyer , using cellular levels of surfactants including LAS and AE , verified a narcotic mechanism of action where the lethal concentration for fish hepatic cells was approximate to fish tissue residues associated with narcosis. The PNEC for the aquatic environment can be determined in several ways based on the quantity of and species and taxa range covered by the available acute, chronic, and mesocosm toxicity data and whether any reliable QSARs have been developed from these data.

For small data sets, the PNEC is estimated using the most toxic standardized e. These surfactants all have large sets of acute and chronic toxicity data, and in many cases mesocosm data; therefore, this simple method is not used. For surfactants with larger data sets of chronic toxicity data, the PNEC can be estimated using several approaches. To apply these approaches, the aquatic toxicity data are first normalized to an environmentally relevant homologue or distribution based on monitoring studies, if available, e.

If there is a large set of chronic toxicity data across a wide range of species and taxa, then the PNEC can be determined using a species sensitivity distribution SSD based on these normalized data using the methods of van de Plaasche et al. SSDs can be used to calculate the concentration at which a specified proportion of species are expected to suffer toxic effects. This concentration is estimated by maximum likelihood assuming a log-logistic or other statistically suitable distribution of these data values, i.

Another approach when chronic toxicity data are available but limited in the species and taxa coverage, as illustrated for LCOH and AS, is to use one or more chronic toxicity QSAR to estimate a toxicity value for the environmentally relevant homologue or distribution. As compared to single-species tests, meso- or microcosm studies involving more than one species are recognized to be a better approximation of environmental reality and therefore, have a higher predictive value.

Therefore, in estimating the PNEC, a smaller application factor is applied to these results if the mesocosm is suitably biologically diverse, contains sensitive flora and fauna, and the exposure is of a chronic time frame Belanger, Typically, this AF is between one and five Solomon et al. Thus, when there are mesocosm studies, these are of greater weight than acute or chronic ecotoxicity data when a weight-of-evidence WoE approach is used to estimate the PNEC for a specific homologue.

These studies are also used in combination with a QSAR based on these data to extrapolate these mesocosm data to an environmental relevant homologue or distribution. Because the surfactants discussed in this paper are generally all HPV chemicals, the detergent industry devoted significant resources to develop mesocosm studies on them.

The objective of this section is to summarize the toxicity understanding from these available data, and how these data are then used to derive the environmental relevant aquatic PNEC. The reader is recommended to refer those documents for specific details of the toxicity data studies. Sufficient measured data exist to quantify the acute and chronic aquatic toxicity of essentially pure alcohols and mixtures of these alcohols to fish, invertebrates, and algae OECD, Because alcohols act by nonpolar narcosis Lipnick et al.

Furthermore, as discussed in Fisk et al. At higher carbon numbers up to C 22 , the measured acute toxicity shows an absence of acute toxicity as evidenced by reported LC 50 values that are greater than the highest test concentration. This is explained by the low water solubility of these LCOHs, which limits their bioavailability, such that an acutely toxic concentration is not achieved Fisk et al.

Effects have also been observed in tests with C 13 and C 14 alcohols but at concentrations that exceeded the solubility of the alcohols; therefore, the observed toxicity may be due to physical effects rather than true toxicity for these two alcohols OECD, The C 14 and C 16 alcohols were not toxic to algae. These results suggest that for alcohols in the range of C 12 —C 14 , there are no acute toxicity effects on fish, invertebrates, and algae driven by the low solubility of these alcohols, although there may be physical effects and that the three types of organisms are about equally sensitive acutely to alcohols of the same chain length Table The acute toxicity data for fish, invertebrates, and algae to multicomponent substances of different carbon chain length alcohols as would be found in commercial products OECD, have also been determined.

These multicomponent substances containing alcohols with carbon numbers in the ranges of C 6 —C 12 , where all the components would be completely dissolved, are acutely toxic at concentrations as expected from the contribution of the individual linear alcohols to the mixture. By contrast for multicomponent substances which contain one or more alcohols with chain lengths greater that C 12 —C 14 , where not all components were fully dissolved, toxicity not only appears to be the result of toxic effects from the soluble portion of the alcohols but also includes toxic effects as a result of physical fouling of the test organism by the longer-chained alcohols.

Chronic aquatic toxicity data for fish and algae OECD, are limited to one or two studies Table Invertebrates, represented by D. There are no data on the acute or chronic toxicity of alcohols to sediment dwelling organisms. As described in OECD , available data suggest that the three taxonomic groups—fish, invertebrates, and algae—are of comparable susceptibility to the individual long-chain aliphatic alcohols, consistent with narcosis structure activity.

Therefore, the database for chronic aquatic effects of single carbon number alcohols from D. Justification for use of the AF of 10 can be found in Belanger et al. However, in the environment, the alcohols will not appear as individual chain lengths but rather as mixtures of chain lengths as evidenced by monitoring studies of WWTP effluents in North America Dyer et al. The most prominent chain length found in those sewage treatment plant effluents across a wide range of treatment types, including lagoons, oxidation ditches, trickling filter, activated sludge and rotating biological contactor, was C 12 —C As described previously, because of the low solubility of LCOH with chain lengths greater than C 15 , no toxic effects will be exerted by these longer chain lengths, and these are thus eliminated from this analysis Belanger et al.

The average chain length based on the range of C 12 —C 15 in the effluent is C The average chain length based on the range of C 12 —C 15 was C 13 after correction was made to the effluent concentrations based on bioavailability corrections for each of the monitored sites based on data in table 4 of Belanger et al. As discussed in Fisk et al. Biotransformation would be expected since alcohols serve as an energy source food through metabolism for a wide range of biota from bacteria to mammals Mudge et al.

INTRODUCTION

The relationship between hydrophobicity and toxicity demonstrated by both the complete acute and chronic toxicity data sets were used as the basis for developing AE chronic QSARs for algae, Daphnia , and fish based on selected studies of specific AE homologues and their K ow. The bulk of the surfactants produced from the detergent alcohols go into household detergents, followed by personal care applications SRI Consulting, a. While exposure can be determined by direct measurement in environmental compartments of interest, such measurements represent the exposure only at a specific time and site situation because chemical usage patterns, wastewater flow rates, wastewater removal efficiencies, and surface water flow rates can vary with time and from location to location. Some possible problems that were anticipated included theft, obstruction of optical devices lenses with foreign objects e. However, toxicity decreased with chain lengths beyond C 14 AS most likely due to solubility constraints. A large number of monitoring studies have been conducted in the United States to determine the removal of LAS in various types of wastewater treatment Table

Branched structures are predicted to have slightly lower BCF values than the corresponding linear alcohols consistent with their lower log K ow. An extensive review and summary of the AE acute and chronic toxicity data with tabular summaries can be found in Belanger et al. The objective of this section is to summarize the toxicity understanding from the available data, and how these data are used to derive the environmentally relevant PNEC for AE.

The reader is recommended to refer those documents for details of the data and any particular studies. EPA estimates that, in , US consumers and businesses discarded televisions, computers, cell phones and hard copy peripherals including printers, scanners, faxes totaling 2. Approximately 25 percent of these electronics were collected for recycling, with the remainder disposed of primarily in landfills, where the precious metals cannot be recovered.

Better data are needed to create a more comprehensive picture of the overall trade flows. Accurate information about the amounts, types of materials and destinations of used electronics exported will provide valuable information for the Federal government, private industry and other stakeholders. A workshop in July in Washington, DC gathered input from stakeholders to and helped chart a path forward.

It presents a methodology for using existing trade data to calculate US exports and lays out challenges and options for future data-gathering efforts. This study involved assessing and mapping flows of electronics, including those exported from the U. View the study here. The strategy provides four overarching goals, action items under each goal, and the projects that will implement each action item. A Update to the National Strategy for Electronics Stewardship , a report detailing key domestic and international accomplishments as a result of the release of the National Strategy for Electronics Stewardship in July These efforts developed the following publications: Solving the E-waste Problem Step: StEP develops scientific papers that help members address e-waste issues within their own organizations and provides global, objective and scientifically-based information that is relevant to addressing the global problem of e-waste.

The repeater retransmitted the data to a receiver module which processed the data and uploaded it to the vendor via a cellular modem. Two piezoelectric products from different vendors had been planned to be tested in Phase 1, but one product experienced problems with vendor support and was therefore removed from the Phase 1 research program. The passive infrared counter that was tested in Phase 2 was part of a combination counter that included a piezoelectric sensor for counting bicyclists. The site used in Arlington was the same test site on the Four Mile Run multiuse trail used in Phase 1.

The site was selected as both piezoelectric counters being tested were already located there, one having been installed by Arlington County in April prior to Phase 1 and the other having been installed during Phase 1. The test sites in the other two cities were selected in consultation with local agency staff and the vendor on the basis of: Two piezoelectric counters were already located there.

The piezoelectric sensor detects bicycles only, the passive infrared sensor detects both pedestrians and bicycles, and the difference in the two counts is the pedestrian count. The Di identified du ew. The therm good field o uired interfa tersection a ment. The radar s io, was buri an access po ation receiv the access p e vendor. Th stallation an ek. The com earby inters t relayed da ss point that.

The primary evaluation criterion was accuracy. However, ease of installation, labor requirements, security, maintenance requirements, software requirements, cost, flexibility of data, and other characteristics were also assessed. Data reducers can also rewind the tape if necessary to check that no pedestrian or bicyclist was missed. This method also had the benefit of allowing a relatively large amount of video to be collected, from which the researchers could select only the time periods with environmental conditions e.

Finally, the method allowed data to be collected at a different time from when it was reduced, which helped spread out the use of research team labor. The process used to generate the ground truth counts is described in a subsequent section. The tube counter was moved from site to site and only installed during a period when a camera was present collecting video for the ground truth counts.

I. INTRODUCTION

Volume 26 contains the works V. I. Lenin wrote between. September and .. scale heroism on the part of worker Bolsheviks in collecting money for remainder translated from In another place they will be put to cleaning latrines. V. I. LENIN. V. I. L E N I N. cOLLEcTED Volume 39 of the Collected Works contains Lenin's Note- .. V. I. LENIN. 30 the remainder (4, million) being in bills. Yet another example: overhauling and cleaning of boilers.

The first occurred after the devices had been calibrated and all parts of the system e. The second occurred later, when warm weather was forecast, to capture higher volumes of bicyclists and pedestrians. The pneumatic tubes were in place during the first ground truth data collection period for each site.

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As described in detail in Chapter 3, correction factors were developed by comparing the ground truth counts to the counts produced by each device that was tested. The researchers selected specific hours from each week of video to develop ground truth counts, corresponding to hours when specific conditions of interest occurred e. Potential installation difficulties include the need for specialized equipment e.

The level of technical assistance required from vendors to install and calibrate the technologies was noted i. Labor requirements include the time and effort needed for installation, ongoing maintenance, data cleaning, and data analysis. Durability and security were evaluated based on both inspection of the hardware to assess the relative durability and resistance to tampering or theft, as well as recording any damage sustained while a device was in place.