HISGAN STAE mess ' lIIIIIIII IIZIIIII IIIIIIIIIIIIIIIIIIIIIIIIIIIIIIII 01565 5040 2. LIBRARY Michigan State Unlversity. This is to certify that the thesis entitled The Role of Biodegradation and Sorption in Determining the Fate of Nonionic Surfactants in the Environment presented by Kathleen A. Wight has been accepted towards fulfillment of the requirements for M. S . degree in Environmental Engineering Major professor Date April1 1997 0-7639 MS U is an Affirmative Action/Equal Opportunity Institution PLACE ll RETURN BOX to remwe We checkout from your record. TO AVOID FINES return on or bdore dete due. DATE DUE DATE DUE DATE DUE Wales-9.1 THE ROLE OF BIODEGRADATION AND SORPI'ION IN DETERMINING THE FATE OF NONIONIC SURFACT ANTS IN THE ENVIRONMENT By Kathleen A. Wight A THESIS Submitted to Michigan State University in partial fulfillment of the requirements for the degree of MASTER OF SCIENCE Department of Environmental Engineering 1997 2m- \“ ill? Ln. 5;. ABSTRACT THE ROLE OF BIODEGRADATION AND SORPI‘ ION IN DETERMINING THE FATE OF NONIONIC SURFACT ANTS IN THE ENVIRONMENT By Kathleen A. Wight Research studies related to nonionic surfactants —- alcohol ethoxylates and alkylphenol polyethoxylates -- are reviewed. Surfactants are defined; unique structural characteristics are identified. The importance of molecular structure, biodegradation kinetics, physical-chemical properties, and environmental conditions in determining behavior and fate are discussed. Possible biodegradation pathways are presented. Data from the most recent environmental studies are highlighted. Environmental levels of surfactants and measurements from toxicity studies are compared. Results indicate that alcohol ethoxylates biodegrade readily to products that are environmentally safe. However, metabolites of the alkylphenol polyethoxylates - alkylphenols, alkylphenol mono- and diethoxlyates, and carboxylated alkylphenol mono- and diethoxylates, sometimes halogenated -- accumulate and are an environmental concern in some areas. The paucity of knowledge about the interrelationship between biodegradation and sorption, and the environmental impact of these derivatives are addressed. To all who have given me wings to soar with the eagles, especially to Dad who taught me about dreaming and to Mom who taught me realism; their undying support provides more strength than they will ever know. iii ACKNOWLEDGMENTS The writing of this paper would not have been possible without the technical assistance and support of the many people named below. I include this section at the risk of inadvertently forgetting to name individuals who have been instrumental in assisting me with any of this paper's many phases. In this event, I apologize to any of you whom I may forget. My omission makes your contribution no less important. I acknowledge Craig Criddle, my adviser, for contributing his expertise, advice, support, enthusiasm, and patience. Craig's wealth of knowledge about organic compounds and microbial processes, combined with his desire to share and ability to articulate his knowledge, have contributed towards an extremely worthwhile learning experience. His criticisms and suggestions have been invaluable in improving the quality of this paper. I acknowledge Bill Punch and Iarry Forney for providing advice, support and encouragement before and throughout the writing process, and for taking the time to review and comment on the first draft. Bill's proposed inclusions and suggestions regarding formatting are much appreciated. Larry's comments and suggestions, a reflection of his own expertise in the area of microbial ecology, have added another dimension to this paper. His responsiveness to my queries, both preceding and during the research for this paper, has been unequaled. iv I acknowledge the many individuals who facilitated my document search efforts by providing access to references related to surfactant research - Jennifer Field, Bob larson, Sharon Grimm, Bruce Dale and Carl lira in the MSU Department of Chemical Engineering, Tom Volkening and his staff of the MSU Engineering library, the staff of the MSU Main library, the staff of the MSU Chemistry library, and the staff of the Oakland University Kresge Library. I acknowledge Kaye Lynn Mazurek and Sachiko Pickler for assisting with the translation of German and Japanese articles, respectively. I acknowledge Linda Steinman for keeping me straight on all the administrative details involved in this effort. Finally, I acknowledge my family -- my husband, Dick Pallister; my son, Doug Pallister; and my daughter, Gail Pallister. Their forbearance, their rhetorical contributions, and their feedback regarding the comprehensibility of figures and graphs have helped immensely during this process. [\TRC RIG! EDIE IRE R! MCDE Firm TABLE OF CONTENTS LIST OF TABLES - - - ......... - viii LIST OF FIGURES - _- - - -- - x LIST OF ABBREVIATIONS ................ _ - _ - -- - xii INTRODUCTION - ...... 1 SURFACTANTS - - - -- -- S Definition/Major Categories ......................................................................... 5 Structure..- ...... _ -- -- - - - - - 5 Detergency .......... -- - - - ................ . _ .......... -- . 9 Nonionic Surfactants .................................................................................... 10 Alcohol ethoxylates -- manufacture .............................................. 1 1 Alkylphenol polyethoxylates -— manufacture ............................. 1 2 ENVIRONMENTAL COMPARTMENTS OF CONCERN - - - -- - 13 Sewer Distribution Systems .......................................................................... 13 Sewage Treatment Plants .............................................................................. 13 Surface Waters and Sediments ..................................................................... 17 Groundwater __ ______ - -_ __ _ -- 17 THE ROLE OF BIODEGRADATION IN ASSESSING ENVIRONMENTAL SAFETY ......... 19 Degradation Kinetics ..................................................................................... 20 Other Factors Affecting Biodegradation .................................................... 26 BIODEGRADATION PATHWAYS - - - - -- - - 31 Alcohol Ethoxylates ....................................................................................... 33 Omega-oxidation of the hydrophobe .............................................. 35 Central fission .................................................................................... 36 Elimination of the ethoxy chain .................................................... 37 Direct hydrolysis ................................................................... 37 Ethylene glycol biodegradation ......................................... 39 Modified hydrolysis .......... - 4O Dehydration-rehydration ................................................... 42 Oxidative hydrolysis - 42 Additional studies _ 43 Alkylphenol Polyethoxylates ..................................................................... 43 PATHWAYS, EXTENT, AND KINETICS OF BIODEGRADATION - GENERALIZATIONS .................... . ......................... 46 vi ENVIRONMENTAL ANALYSIS AND SAFETY ASSESSMENT ....................................... 49 Alcohol Ethoxylates -- - S 1 Sewer distribution systems - - - 51 Sewage treatment plants - -- - -- - - - 53 Primary treatment - ........ - 53 Secondary treatment-- - - -- - - 56 Surface waters -- - - -- - - ...... S7 Groundwater -- - - - - - _- - - -- 59 Summary -- 60 Alkylphenol Polyethoxylates ........................................................................ 61 Sewer distribution systems - ..... - -- -- - -- -- 62 Sewage treatment plants---- - - - 63 Primary treatment - -- - - 67 Secondary treatment ----- - - - -------- 68 Sludge digestion - - - . - - - 77 World data - STPs --------- - - ’ 80 Surface waters ...................................................................................... 83 Water column ---------- - - -_ 83 Water column and sediment - - -------- 89 World data- Surface waters - -- -- 91 Groundwater - -- - - 98 Summary ............ - - - - - - - 106 Halogenated Derivatives - -- -- - ------ - - - - 109 Toxicity - - - - -- - 112 CONCLUSIONS and RECOMMENDATIONS - -- -- -- 122 APPENDICES - - - ...... 125 APPENDDI A - SEWAGE TREATMENT PLANTS - - -- 125 APPENDDt B - GLOSSARY OF TERMS - - - 128 APPENDDI C - ESTIMATION OF Kow USING FRAGMENT CONSTANTS ...... 131 APPENDIX D -- N ONIONIC SURPACT ANT BIODEGRADATION PATHWAYS . 132 APPENDD( E— TOXICITY OF NONIONIC SURFACTANTS TO AQUATIC ORGANISMS 135 REFERENCES ................................................................................................................ 138 vii la la; Ia: LIST OF TABLES Table 1 - Worldwide Consumption of Surfactants - - 2 Table 2 - Major Synthetic Detergent Surfactants 8 Table 3 - Activated Sludge Treatment Plant Retention Times 16 Table 4 - Concentrations of Suspended Solids in Selected Media - 16 Table 5 - Factors Affecting Extent and Rate of Biodegradation of Organic Compounds 26 Table 6 - Nonionic Surfactant Physical-Chemical Properties 27 Table 7 - Comparison of Pure and Mixed Aqueous Solubilities of NPE1.4 ......... 29 Table 8 - Organisms Capable of Growing on Nonionic Surfactants or PEG ........ 32 Table 9 - Environmental Analyses in Sewer Distribution Systems .................... 52 Table 10- Environmental Analyses in Sewage Treatment Plants and Receiving Waters (ug/ L) - - - ...... 54 Table 1 1 - liquid/Solid Phase Fractions of Raw and Settled Influent ................ 56 Table 12 - NPE Molar Flows in Ztirich-Glatt STP (moi/day (96 of Total)) ............ 64 Table 13 - 96 Distribution of Nonylphenolic Compounds (Molar-Based) ........... 66 Table 14 - Range of Product Elimination Efficiencies (Aqueous Phase) - Secondary Treatment 70 Table 15 - Temperature Effects on the Extent of APE Removal - 7 1 Table 16 - Removal of NPE1-13 in Two North Carolina STPs (ug/ L) ................... 73 Table 17 - Removal of NPE1-13 in Seven U. S. STPs (pg/L) 74 Table 18 - Variations in Oligomer Behavior During Secondary Treatment ...... 77 Table 19 - Relationship of Concentration of Suspended Solids to Fraction of NP Partitioning to Solids - - - 78 Table 20 - Concentration of NPE and Its Metabolites in STPs (ug/ L) ................. 81 viii la la la Table 21 - Inflow and Outflow of NPEs in the Glatt River 85 Table 22 - Environmental Analyses of N PBS in River Water and Sediment - by Major Category -- 86 Table 23 - Environmental Analyses of NPcs in Fresh Water and Sediment ...... 92 Table 24- Analyses of NPcs in Marine Environments -- Aqueous/Solid Partitioning -- 95 Table 25 - NPc Elimination in Groundwater Underlying the Glatt River ........ 100 Table 26 - Fluctuations in NPE Product Concentrations -- High Readings from the Glatt River and Underlying Aquifer (ug/ L) 103 Table 27 - NPc Concentrations in Groundwater in Zagreb, Croatia (pg/ L) ....... 105 Table 28 - Concentrations of Halogenated OPEs and OPECs Measured in Secondary Effluents in Palo Alto, California -- 11 1 Table 29- Toxicity of Nonionic Surfactants to Aquatic Organisms - Summary (pg/ L) ....... 1 14 Table 30 - Bioconcentration Factors of Nonylphenol Ethoxylate Derivatives in Fresh Water Animals and Plants 1 19 Table 31 - Bioconcentration Factors of Nonylphenol in Marine Organisms 120 Table 32 — Toxicity of Nonionic Surfactants to Aquatic Organisms ................... 135 ix LIST OF FIGURES Figure 1 - Examples of hydrophobe/hydrophile arrangements in detergent surfactants .................................................................................................. 6 Figure 2 - Surfactants clean by wetting, dispersing and emulsifying .............. 9 Figure 3 - Environmental compartments affected by detergent surfactants .. 14 Figure 4- Generalized flow diagram of an activated sludge treatment plant -- - 15 Figure 5 - Major proposed biodegradation pathways for AEs - 34 Figure 6 - Proposed biodegradative «lo-oxidation pathway for AEs ..................... 36 Figure 7 - Proposed central fission pathways for AEs ......................................... 38 Figure 8 - Proposed biodegradation pathways for the ethoxy chain of ABS, APEs and PEGs 41 Figure 9 - Proposed biodegradation pathways for APEs ...................................... 44 Figure 10 - Relative ease of biodegrading ABS and APEs ..................................... 46 Figure 11- Concentrations of ABS in raw influents, and primary and secondary effluents ..... - ------------------ - SS Figure 12- A greater proportion of AES is in the aqueous phase followrng 1° treatment-- - 56 Figure 13 - Distribution of individual NPE“ oligomers in raw sewage, primary effluent and secondary effluent of the sewage treatment plant Uster and a commercial mixture Marlophen 810 -- - -- - - - - - -- -- 63 Figure 14 - NPE molar flows in the Ziirich-Glatt STP ............................................ 65 Figure 15 - Relative importance of biodegradation and sorption as removal (I) and accumulation (1) mechanisms during water treatment - by major category .............................................................................. 67 Figure 16 - Concentrations of NPE and metabolites in secondary effluents 82 Figure 17 - Distribution of NPE compounds in secondary effluents and Glatt River water ..................................................................................... 85 X Figure 18 - Longitudinal profiles (concentration and mass flows) of measured NPE derivatives and total measured NPc -- 88 Figure 19 - Longitudinal concentration profiles of NP, NPE1, and NPEz in sediment 90 Figure 20 - NPE1-20 marine environment concentrations, from untreated sewer waters discharged to Sibenik Harbor, Croatia: (a) in dissolved and particulate fractions based on water sample volume, and (b) based on dry weight of suspended solids .............. 96 Figure 21 - Concentrations of NPEp in the vertical profile of Sibenik Harbor, Croatia 97 Figure 22 - Longitudinal profile of average concentrations measured in Glatt River and underlying aquifer, at 0.5 m below the water table 100 Figure 23 - Seasonal variations in NP, NPE1, NPEz concentrations, measured in river water and groundwater, 0.5 m below water table 102 Figure 24 - NP and NPEz concentrations fluctuate in aquifer during winter months - -- - - - - - - 104 Figure 25 - Proposed APE pathways in an STP with chlorination ..................... 110 xi LIST OF ABBREVIATIONS primary secondary alcohol ethoxylate (p = no. of ethoxy units) alcohol ethoxy sulfate alkylphenol alkylphenolic compounds alkylphenolic compounds alkylphenol ethoxylate (p = no. of ethoxy units) carboxylated alkylphenol ethoxylate (p = no. of ethoxy units) alkyl sulfate activated sludge bioconcentration factor biodegradation half-life biochemical oxygen demand designation for branched alcohol ethoxylate with 'm' carbon groups and 'p' ethoxy units designation for alcohol ethoxylate with 'm' carbon groups and 'p' ethoxy units 2,6-dichlorophenolindophenol dodecylphenol ethoxylate diethanolester dimethylammonium chloride (cationic surfactant) diethylene glycol dinonylphenol ethoxylate dissolved organic carbon decylphenol ethoxylate effect concentration at which a 50% reduction in growth rate is observed ethylene glycol ethoxy group (-CHz-CHz-O-) xii FAB/ MS GA GC/IVBD HPLC HRT I & I I R LAS LC LC/MS LCso 10K: MCRT MLSS MLVSS NAD NADH N MR NOE: NP NP-HPLC N Pc N PEp NPEpC OBS OPEp OPEpC PAE PB-LC / MS PEG Ppb PPm PDQ fast atom bombardment with mass spectrometry glucose amide (nonionic surfactant) gas chromatography with mass selection detection hydrocarbon high pressure liquid chromatography hydraulic retention time institutional and industrial infrared linear alkylbenzene sulfonate liquid chromatography liquid chromatography with mass spectrometry lethal concentration at which > 50% of tested organisms die lowest observed effect concentration maximum accepted toxicant concentration mean cell-residence time monoethylene glycol mixed liquor suspended solids mixed liquor volatile suspended solids nicotinamide adenine dinucleotide (oxidized cofactor) reduced nicotinamide adenine dinucleotide (reduced cofactor) nuclear magnetic resonance no observed effect concentration nonylphenol normal-phased high pressure liquid chromatography nonylphenolic compounds (NPE1-20, NPE1&2C, NP) nonylphenol ethoxylate (p = no. of ethoxy units) carboxylated nonylphenol ethoxylate (p = no. of ethoxy units) optical emission spectroscopy octylphenol ethoxylate (p = no. of ethoxy units) carboxylated octylphenol ethoxylate (p = no. of ethoxy units) primary alcohol ethoxylate particle beam-liquid chromatography with mass spectrometry polyethylene glycol parts per billion parts per million parts per quadrillion xiii ”x Ppt QSAR RP-HPLC RS M RT SAE sec-CmEp SIM-GC/ MS STP TBS TF TLC TLC/ RAD UV WWTP parts per trillion quantitative structure-activity relationship reversed-phase high pressure liquid chromatography resuspended solid materials residence time secondary alcohol ethoxylate secondary alcohol ethoxylate with 'm' carbon groups and 'p' ethoxy units selected ion monitoring gas chromatography with mass spectrometry sewage treatment plant tetrapropylenebenzene sulfonate trickling filter thin layer chromatography thin layer chromatography with radio scanner ultraviolet wastewater treatment plant xiv Int: mic dere dew the "I Lu tar Introduction For close to half a century synthetic detergents have been the subject of biodegradation and environmental fate research studies. Not long after the initiation of their mass production in late 1946, the effect of synthetic detergents in the aeration tanks of activated sludge sewage treatment plants was obvious, as layers of sticky, bulky, smelly suds up to eight feet thick developed in some cases (1). Plant operators measured less dissolved oxygen in the wastewater (2), thus more organics in the plant effluent, and observed less settling and filtration of solid sewage material (1). The general public was alerted to this situation when foam appeared and remained in receiving natural waters downstream from outlets of wastewater treatment plants (1 -3) . Concern about the long term effects of the new synthetic detergents in our sewage treatment plants, surface waters, and ultimately, our drinking water increased as their use increased. Soon studies identified surfactants, specifically the highly branched, biorefractory tetrapropylenebenzene sulfonates (TBSs), as the detergent components that were causing these adverse conditions. Pressured by consumers and water treatment personnel, surfactant producers replaced TBS with linear alkylbenzene sulfonate (LAS), a more biodegradable surfactant, by mid-1965 (1). Since then, other surfactants -- nonionic and cationic -- have come increasingly into use. Coincident with efforts to find readily biodegradable components, the development of methods to test biodegradability and to determine the environmental fate and safety of detergent components, have continued to be the focus of research studies in the detergent industry. Surfactants, comprising approximately 25% by weight of detergent formulations (4), have received most of the attention. Surfactants are noted for their success in a wide range of products which have diverse uses in household, personal care, It) 2 institutional, and industrial markets (5 - 7). Total worldwide surfactant consumption during 1995 was approximately 8.4 million metric tons (Table 1). The North American share of this total was about 3096, 2.5 million metric tons. More than half of the 1995 world total, approximately 4.9 million metric tons, was used in consumer products in the form of detergents, dishwashing and cleaning agents, cosmetics and toiletries (8). Most of the remainder was used in a variety of industrial applications, such as lubricants, pastes, waxes, plastics, cements, plaster, and hydraulic fluids (5, 8). Surfactant consumption continues to grow, particularly in Asia and the third world countries (Table 1). Table 1. Worldwide Consumption of Surfactants‘ 1995 2005 Change A. By Region (106 metric tons) (106 metric tons) (96) North America 2.5 2.8 12 W. Europe 1.8 2.2 22 Asia 2.6 4.0 54 Other Regions L5 L3 53 Total 8.4 11.3 35 B. By End Use Household 4.3 5.8 35 Personal Care 0.6 0.9 50 I 8: l/lndustrial 1.; 4_.6 31 Total 8.4 1 1.3 35 ‘ Ref. (8); soap not included. The sense of urgency in understanding the fate of surfactants is attributed to; 1) their widespread use and release by households, industries and institutions, in large quantities (U.S. consumption of household and personal care products was over 6.8 million metric tons in 1993 (8)); 2) knowledge gained through biodegradation and environmental fate studies that the metabolites of some surfactants, particularly the alkylphenol polyethoxylates (APEs), appear to persist and 3 accumulate in the environment (9-1 7), sometimes as halogenated derivatives (18-23); 3) increased awareness that some of the persistent APE metabolites are more toxic or more likely to bioaccumulate than their parent compounds (24-27); and 4) concern that some persistent APE metabolites, acting as estrogen mimics, may cause irreversible ecotoxicological damage in as yet unknown ways (28-34). Predicting the environmental fate of any chemical relies upon gaining knowledge about the biological, chemical, and physical processes that determine its transformation (change in chemical structure) and disposition (no structural change) in the environment. A myriad of occurrences are possible - deposition, volatilization, precipitation vs. dissolution, vertical and horizontal mixing, diffusion, sorption, sedimentation, bioturbation, resuspension, and photochemical, chemical, and biological reactions (35). Designed to be applied in water, surfactants are highly soluble and negligibly volatile. Thus biological reactions and sorption followed by sedimentation play major roles in determining fate in the environmental compartments in which detergent components are found (6, 15, 36, 37). Bioturbation, resuspension, volatilization, and photochemical reactions may be minor processes in the aquatic environment. This paper reviews the results of recent research related to nonionic surfactants. Surfactants are described, with emphasis on the major ones found in synthetic detergents. Reasons for the evolving significance of nonionic surfactants are enumerated. The environmental compartments that are affected by the use of detergent surfactants are identified. The importance of biodegradability tests and biodegradation kinetics in determining 4 environmental acceptability is discussed, as is the importance of sorptive processes. Possible biodegradation pathways are presented. The most recent findings from environmental studies are highlighted, primarily those not discussed in previous comprehensive reviews (6, 27). A summary of data measured in environmental analyses is provided. Emphasis is placed on the role of microbial degradation, while also considering the effecrs of sorption. A discussion about the results of studies that have detected halogenated derivatives of alkylphenol ethoxylates in treated wastewater, surface waters, and drinking water is included. Integral to this review is a discussion of the manner in which molecular structure, physical-chemical properties, and environmental conditions determine surfactant fate. Concerns about the persistence, accumulation, and toxicity of surfactants and their metabolites are addressed. Surl- Defi ”.1 (I) b ‘ Sll.’ 4 Surfactants Definition/ Major Categories A surfactant, short for "surface active agent", functions as a coupling agent between two surfaces, bringing them closer together because of the way the molecules orient themselves at the interface of the two surfaces (1). Surfactant molecules tend to concentrate at air-water, solid-water, or non- aqueous liquid-water interfaces (6). This tendency exists because structurally, surfactants are amphiphiles, with polar (hydrophilic) and non-polar (hydrophobic) moieties. The molecules orient themselves so that the hydrophilic (water-attracting) ends are in the water and the hydrophobic (water-avoiding) ends are positioned as far away as possible in the opposite direction (1, 6). Differences in the ionic properties (electrical charge) of the hydrophilic groups in water are the basis for classifying surfactants into one of four large groups -- nonionic, anionic, cationic, or amphoteric (zwitterionic). Within these groups surfactants can be further classified based on the means of linkage between the hydrophilic and hydrophobic moieties. The hydrophobic component may take one of several forms, but does not affect the classification of the surfactant into the four large groups named above (6, 7). Structure Hydrophobic groups are derived from fatty acids, olefms, alcohols, alkylbenzenes, alkylphenols, paraffins, or polyoxypropylenes. A hydro- carbon of between eight and twenty carbon groups is usually seen in the hydrophobic moiety of surfactants. The derivatization process may produce linear or branched alkyl groups. The benzene or phenol ring can be linked anywhere along the alkyl chain; ring substituents may be present in the para-, ortho-, or meta-position. A large number of various structures is 6 possible from the many available combinations. Additionally, the character of some hydrophobes (e.g. alcohols and alkylphenols) may be modified by adding propylene oxide. The longer the polyoxypropylene polymer, the more hydrophobic the compound. All hydrophiles possess polar groups that make them soluble in water. Hydrophilic groups may or may not ionize in aqueous solution. The many uses of surfactants derive naturally from the permutations and combinations that are possible with the large selection of hydrophobic and hydrophilic groups, and the ways of linking them together (6). Figure 1 provides examples of configurations for the three surfactant categories that are most often used in detergents. —O>O——-OO O© I —<:> @ + .1... o Fugue 1. Examples of hydrophobe/hydrophile arrangements in detergent surfactants. Adapted from (5). 7 Nonionic surfactants do not dissociate into ions in aqueous solution. Anionic and cationic surfactants dissociate into negative and positive ionic fractions, respectively. The negative ion carries the surface active properties of the anionic surfactant and the positive ion carries the surface active properties of the cationic surfactant. A nonionic group commonly seen is polyoxyethylene ((-O-CH2-CH2)n-OH). Anionic groups are sulfonate (-SO3'), sulfate (-OSO3‘), and carboxylate (-COz'). Quaternary ammonium (-R4N+) is an example of a cationic group. Amphoteric surfactants are molecules that carry both a positive and a negative charge in aqueous solution. Whether they have anionic or cationic properties is dependent on medium pH and composition (5, 6). Amphoteric surfactants are not generally used in detergents (4). Example molecular structures of the five surfactants (2 nonionic, 3 anionic) used most worldwide are shown in Table 2. In this paper, CmEp denotes a primary linear AE molecule that consists of an average of m alkyl carbons and p EO units. Sec-CmEp or br-CmEp is used if the alkyl moiety is secondary or highly branched, respectively. APEp (p=1-20) denotes an alkylphenol polyethoxylate that has p ethoxy groups. Oxidation of the terminal ethoxy unit during biodegradation may result in a carboxylated derivative. APEpC denotes a carboxylated alkylphenol polyethoxylate that has p-1 ethoxy groups terminated with an acetic acid group. AP refers to an alkylphenol (no substituents). The prefixes OP-, NP-, DP-, and DDP- are used in place of AP- to define octyl-, nonyl-, decyl-, and dodecylphenol, respectively. 8 Table 2. Major Synthetic Detergent Surfactants m,n - No. of carbon gps in hydrophobe; Common name Acronym Chemical structure p - No. of ethoxylate gps. in hydroghile; ‘ Nonionic surfactants Alcohol ethoxylates AE Primary: -primary PAE Rm-O-(CHz-CHz-O )p-H m - 8 - 18;1 Secondary: -secondary SAE C(Rm)(Rn)—O-(CHz-CHz-O )p“ fills 214-91 +n s 18;‘ p-l-ZS2 paw - 6 ’ 1 I 2 Alkylphenol APE R... R... is usually branched, m - 8 - 12;1 polyethoxylates with 8 or 9 carbon groups. , p-l-ZO3 I Pave ' 9 ' IO 1 0—(CHz-CHz-0)p-H R... - CHz-CH-CH z-CH-CHz-CH- I I l CH3 CH3 CH3 Anionic surfactants Linear alkylbenzene LAS 10 s m+1+n s 15 4'5 sulfonates Alcohol ethoxy sulfates AES Rn-O-(CHg-G-lg-O)p-803Na m - 12-15 6 p - 3-7 ‘ Pavg ' 2'4 6 Alkyl sulfates AS Rm-O-803Na m - 8-18 7 Adapted from (5). Other references: ‘ (27); Z (38); 3 (39-41); 4 (6'); 1' (42); 5 (43); 7 (44). De tergency Surfactants function as cleaners because as they concentrate at inter- faces, they reduce the surface tension of water. At the air-water interface foam is produced, as seen in treatment plants and natural waters subsequent to their introduction (1, 6). Surface active agents may possess other properties -- wetting, dispersing, and emulsifying - to various degrees. In a cleaning agent, or detergent, all three properties exist to approximately equal degrees (1). The wetting property lowers the solid-water interfacial tension, thus displacing air from solid surfaces and allowing penetration of a water solution into fabric where it can reach soiled material. At the solid-water and non- aqueous liquid-water interfaces, the dispersing property separates agglomerated particles of oil and dirt; the emulsifying property links water molecules with these particles, thus holding them in suspension (Figure 2). Soil removal occurs because the hydrophobic ends of the surfactant molecule surround soil particles as they are attracted to them. Simultaneously, the hydrophilic ends are attracted back into the wash water, lifting the soil particles away from the fabric. This action, together with the mechanical agitation of a washing machine, provides effective soil removal and suspension while preventing redepositon (1, 5, 6). WW 5) A"‘“'°' . . interface ‘ ‘ \03 e e O\ . \ ET H‘o’H HRH “pf” 'Q ~ ‘ 4 '. t 0’ I _,.Of;‘0‘_ .s Dirt andoil H ,H Q 9 H H O )3, ‘q 7) ‘ ‘ partlcles W“ ' " - -- TI 5' Solid-water C) Surfactant O O ‘0 6 interface mecules O'OHB'H ‘90\ {db 6 ’64?“ Water 0" O— o ‘9 I H36 H H H molecule Figure 2. Surfactants clean by wetting, dispersing and emulsifying. 10 N onionic Surfactants Worldwide consumption of the most used nonionic surfactants - alcohol ethoxylates (ABS) and alkylphenol polyethoxylates (APEs) - was close to 1.4 million metric tons during 1995. The U. S. share of this was 227 thousand metric tons, over 1696. Though anionics enjoyed about 7596 of market share in 1995, several developments in the synthetic detergent market, some interrelated, are eroding the anionic surfactant market share in favor of nonionic surfactants: 1) 2) 3) 4) S) 6) nonionics are superior to anionics in cleaning synthetic fibers (3); nonionics clean better in cold water, which is preferred over hot water both for cleaning man-made fibers (3) and for conserving energy (45); the nonionics, especially the linear primary alcohol ethoxylates (PAEs), continue to gain favor over the linear alkylbenzene sulfonates (LASs) because of their rapid biodegradability (45 ); the gradual shift from dry powder to liquid detergents (45, 46) favors increased usage of nonionic surfactants (45, 47); pushed by speculation that an upcoming U. S. Department of Energy directive will require more energy efficient, horizontal axis washing machines that will use 75% less water (8, 48), detergent manufacturers look to nonionics to satisfy requirements for a surfactant that will perform well in cold water with minimal foaming and superb rinsing capabilities (6, 45); and detergent formulations, driven by cost and therefore likely to change according to fluctuating prices of raw materials, periodically favor the alcohol moiety used in alcohol ethoxylates over the alkylbenzene moiety used in LAS (3, 48). 11 During 1995 worldwide consumption of AB and APEs was 726 and 635 thousand metric tons, respectively (8). The largest market for the ABS, mainly linear PAEs, is household laundry detergents and cleaning products (2 7, 49). Additionally, a large share of linear PAEs in the United States provides the alcohol ethoxy moiety of alcohol ethoxy sulfates (2 7, 50). Secondary and branched AEs, as well as the linear PAEs, are used in other specialized household cleaners, personal care products, and institutional and industrial (1 8: I) cleaners (2 7, 49). The APE's primary market in the U. S. is the industrial processing sector (55%). The remaining 4596 is split between institutional cleaners (3096) and household and personal care products (1596) (27). Applications for 'both ABS and APEs exist in agriculture (5, 18), pulp and paper (10), textile and tanning (20), and other process industries (27). The AEs are gaining a stronger position in the nonionic market because concern about the persistence, accumulation, and toxicity of APE metabolites is forcing replacement of APEs with A55 in the I & 1, household, and personal care market sectors (45). Al -- f The manufacture of alcohol ethoxylates requires the reaction of alcohols with ethylene oxide. The alcohols, usually comprised of 8-18 carbon groups, are most often linear primary, but may also be linear secondary or branched (27). linear alcohols that originate from vegetable or animal sources (oleochemicals) have an even number of 12 - 18 carbon groups (38). The number of carbon atoms (usually between 11 and 15) in those that are derived from crude oil (primarily ethylene) or natural gas (petrochemicals) can be even or odd (27, 38, 51). The number of ethoxy units found in the ABs typically used in detergents varies from one to between 15 and 25, with an average of 6 to 11 units (38) (Table 2). 12 Biodegradability is not an issue in choosing between raw materials (51). Results of life cycle inventories indicate that both sources - oleochemicals and petrochemicals - have environmental trade-offs (5 2). Therefore, cost of production and availability, not biodegradability or other environmental considerations (e.g.: air emissions, waste byproducts, consumed energY), are determining factors (45, 5 3). The manufacture of alkylphenol polyethoxylates requires a base- catalyzed reaction of alkylphenols with ethylene oxide. The alkylphenol is produced by reacting branched olefins with phenol (acid-catalyzed). The alkyl group may exist in a variety of isomeric configurations. More than 90% of the alkylphenol products exist in the para-isomer form with less than 10% as ortho-isomers: I'D para-alkylphenol ortho-alkylphenol . Joined to the alkylphenol via an ether linkage is a hydrophilic chain with 1 to 100 ethoxy (150) units (27). In spite of concerns about their environmental persistence, accumulation, and toxicity, use of the APEs continues because they are excellent performers with relatively low production costs (5 ). Octyl- , nonyl— , and dodecylphenols are found most frequently in commercially available APEs (3, 5). Use of 4-nonylphenol polyethoxylates is predominant in detergents (5, 27). Propylene trirner is the alkene used in the manufacture of the nonylphenol polyethoxylates (5, 2 7, 5 4); the hydrophile commonly consists of between 1 and 20 E0 units (39, 40, 41), averaging 9 to 10 units per molecule (27, 40) (Table 2). Trace amounts of 2—nony1-, 4-octyl-, and 4-decylphenol may be present as impurities in commercial 4-nonylphenol (13, 55) . Environmental Compartments of Concern Detergent surfactants, used and disposed of in wash water, or their metabolic products are found in sewage distribution systems, sewage treatment plants (STPs), surface waters, sediments, and groundwater. Figure 3 provides a pictorial diagram of the environmental compartments affected by the release of surfactants. Each of the major compartments is discussed briefly. Sewer Distribution Systems The safety assessment of surfactants begins with analysis of behavior in sewage distribution systems. Sanitary sewers carry domestic and industrial wastewater to treatment plants for removal of solids and organics prior to release to surface waters. Flows from residences, commercial buildings, institutions, recreational facilities, and industries (56) are channeled to interceptor and/ or main sewers by way of collecting sewers before reaching the wastewater treatment plant. Sewer systems are usually designed to carry flows at velocities between 0.6 and 3 meters per second. Depending on the size of the area being served and distance from point of entry to treatment plant, minutes to hours may pass during a waste material's transit through a sewer distribution system (5 7). The extent to which sewers contribute to removal from the water phase is dependent on percent solids (affects sorption) and residence time (affects biodegradation). Sewage Treatment Plants Most sewage treatment facilities are designed to provide secondary as well as primary treatment. Some also offer tertiary, or advanced, treatment. Primary treatment removes settleable or floating solids physically. Secondary treatment removes dissolved organics biologically or chemically; activated sludge is the biological system used most worldwide. Advanced treatment combines operations and processes to remove other constituents, such as 13 \ t Unsaturated Zone 14 5”” I ' Resuspension Sedimentation 1' fi fi fi ‘1 l A S <- cm \Sediment Boundary Layer W I I <10cm (oxic region) 10's cm \ - __Water table :Infiltlation. *--....___v__v v ~ AQUIFER "e—i Frame 3. Environmental compartments affected by detergent surfactants. Adapted from Ref. (58, 5 9). 15 nutrients or priority pollutants, that are not removed during secondary treatment. Figure 4 is a generalized flow chart of the activated sludge sewage treatment plant seen commonly in the United States (5 6). Biofilters would replace the aeration tank in a flow chart for a trickling filter plant. An important aspect of the sewage treatment process is sludge digestion and disposal. Stabilization of sludge occurs during sludge digestion, which further decomposes organic and inorganic matter. The process is commonly designed to occur anaerobically, although aerobic digestion is also used, usually for smaller wastewater treatment facilities. Ultimately secondary effluents are disposed of in surface waters. Though there are many ways to dispose of digested sludge (5 6), the effects of land application only will be discussed in this paper. Appendix A provides a more detailed description of the major processes in sewage treatment facilities. Approximate times for the various waste processing segments are provided in Table 3. Times will vary under different load, temperature, and other environmental conditions (56). Typical ranges of suspended solids concentrations, found in a variety of environmental matrices, are provided in Table 4. lnlluent Screens &Ior ‘ em I bet Aeration tmk ' I y . - Recycled dodge Screenlnge - ..................... g - _L_ Belt ‘ ...-.-..-..... 5'" press K 1 Wm ——-> W smm~------> W * Flgtre 4. Generalized flow diagram of an activated sludge treatment plant. Adapted from Ref. (56). l6 Tfile 3. Activated Sludge Treatment Plant Retention Times Process Time Ra_nge Primgry — Primary settling tank 1.5-2.5 hrs Secondary - Aeration tank (hydraulic) 3-8 hrs - Aeration tank (cell) 5-15 days - Clarifier 3-4 days Contact Tank 15-45 mins Anaerobic Digester - Standard 30-90 days - High rate 10-20 days Ref. (56). Table 4. Concentrations of Suspended Solids in Selected Media Concentration Range Medium (mL/L) Untreated domestic wastewater 100 - 3501.2 Aeration tank (MLVSS) 1,200 - 6,4001-3 Primary and waste activated sludge 30,000 - 80,000"4 Secondary sludge (w/primary settling) 5,000 - 15,000“4 Septage (cess pools and septic tanks) 2,000 - 100,000‘ '5 Surface water 0.5 - 600G ‘ Ref. (56). 2 Table 3-16, p. 109. 3 0.80 * MLSS (56, p. 390 and Table 105, p. 550). 4 Table 12-8, p. 774. 5 Table 14-22, p. 1092. 5 Ref. (60). 17 Surface Waters and Sediments Surface waters of interest in this paper include rivers, tributaries, lakes, and estuaries. Secondary effluents, usually discharged into a river or its tributaries, are the predominant source of surfactants in surface water. Behavior and fate studies focus on surfactant concentration levels (in effluent, surface water and sediment samples), stream flow, dilution factors, colloidal material, suspended solids concentrations, bio-transformation/ sorption rates, sedimentation rates, bioconcentration factors, and toxicity measurements. The upper limit of 600 mg/ L given for the concentration of suspended solids in surface waters (Table 4) may be low in some instances. According to (61), solids concentrations in natural water systems range between 10 and 10,000 mg/ L, where 10 mg/ L represents the lower limit of suspended solids concentrations in rivers and estuaries, and 10,000 mg/ L represents the lower limit of solids concentrations found in sediment beds. Therefore it is conceivable that turbulence caused by high winds or heavy flow in a natural water system could increase suspended solids to levels above 600 mg/l.. Groundwater Although surface waters, as the sink for secondary effluents discharged from STPs, have received most of the attention in environmental fate studies, they account for less than 5% of the earth's total freshwater supply. Approximately 95% of the freshwater used for domestic, commercial and industrial purposes is obtained from aquifers. Due to this distribution of freshwater supplies, groundwater is ever more important as a resource to meet increasing demands worldwide (5 9). This explains the concern about groundwater contamination, and recent investigations that have examined the behavior and fate of pollutants and microorganisms in groundwater (62- 73). 18 Groundwater regions are subject to long-term contamination because their characteristics -- lesser amounts of mixing and dilution, low temperatures, and reduced microbial diversity and population sizes - are not favorable to high rates of biodegradation (64). Behavior and fate predictions about movement (advection and dispersion), sorption (retardation), biotransformation, and resulting concentration levels through subsurface regions are complicated by the heterogeneous nature of aquifer material. The complex relationship between hydro-geological processes and contaminant behavior and fate dictates an interdisciplinary approach to groundwater research. Sampling must be much more extensive than in surface waters to ensure that contaminant plumes are properly identified. Subsurface residence times can vary from a relatively short period of several hours (62) to a very prolonged period of up to thousands of years (5 9). Studies begin by making simplifying assumptions that address only the major mechanisms at work in the groundwater environment. Given appropriate residence times and environmental conditions favorable to microbes, sorption and biological transformation are again the major removal mechanisms of surfactants in these regions. The Role of Biodegradation In Assessing Environmental Safety Environmental safety is assessed by evaluating data that measures ecological hazard and exposure (RISK - HAZARD * EXPOSURE). Results from toxicology studies, discussed in a later section, provide information for the HAZARD part of the risk assessment equation. Results from biodegradation and environmental fate studies provide information about the EXPOSURE part of the equation. Acceptance of the risk assessment is more probable as certainty increases that study results reflect what happens in the environment. A variety of approaches, test methods, and analysis techniques for determining biodegradation pathways, and extent and rate of biodegradation have evolved from concern about surfactant safety. Determination of biodegradation pathways is very complex and must be carried out in the laboratory under closely controlled conditions (2). Determination of the extent (biodegradability) and rate (kinetics) of biodegradation is conducted under both laboratory and field conditions (74-78). Environmental fate studies consider other factors, such as sorption, that may affect a compound's ultimate disposition. A host of terms related to chemical structure, biodegradability, water treatment, and affected environmental compartments has evolved from biodegradation, environmental fate, and toxicology studies. A glossary of these terms is provided in Appendix B. To the extent that biodegradation is a major removal mechanism in the environment, thus reducing exposure, the assessment of biodegradability through standardized tests continues to be of primary importance when considering the introduction of all new component molecules to detergent formulations (79-84). Guidelines established by the Organization for Economic Cooperation and Development (OECD) suggest a tiered approach (three test levels), that acknowledges the inadequacy of laboratory testing to evaluate 19 20 ecological impact. The OECD guidelines are used widely by the detergent industry. The first level test is stringently designed and least resembles environmental conditions. Though the test is carried out under aerobic conditions, the test substrate is offered as the only source of carbon; the test substrate concentrations are higher and the active biomass concentrations are lower than found in the environment. Compounds that achieve ultimate biodegradation to acceptable levels within a specified time period during the first level screening test are classified as readily biodegradable. Conditions for the next two test levels are adjusted so that they are progressively more environmentally relevant. Compounds that meet the second level test requirements are classified as inherently biodegradable. Conditions in the third and final test in the series simulate the real world as closely as possible. This test assesses the rate as well as the extent of biodegradation (82, 83). Degradation Kinetics Microorganisms fulfill their requirements for energy and growth by catabolizing inorganic and organic compounds. Transformations occur in the presence of enzymes that organisms produce to fulfill their own metabolic needs (85). The rate of transformations are dependent on the level of requisite enzymatic activity, which is a result of microbial growth (86). Mathematical expressions describe the rate at which microorganisms grow, the rate at which breakdown occurs (substrate utilization), and the relationship between the two (85). 0 Growth rate of microorganisms: u=um S/(S+Ks) -b [1] um=1> Sm", - S is the primary substrate, and net growth occurs; 2) S << Sm,n - Sis a secondary substrate; a) the substrate concentration is too low to support growth, but in the presence of a primary substrate that supports growth, the organism may still be capable of consuming (transforming) the secondary substrate; b) there may be several secondary substrates - 81, 82, $3, ..., Sn - all at very low concentrations; versatile organisms might utilize many of the low concentration substrates simultaneously, such that they generate enough energy to satisfy both maintenance and net growth requirements; 3) co—metabolism occurs when an organism does not obtain energy from a transformation that is catalyzed by enzymes the organism produces; a molecule is transformed fortuitously. Further examination of the substrate utilization equation [3] yields two different relationships when the substrate concentration S is extremely large or small relative to the half-velocity coefficient [(5. dS/dts-anwhenS» Ks [6] dS/dt = - (km/1(5) X S when 8 << Ks [7] k' = k,,,/I(s [8] Note that substrate utilization is first order (km is the rate constant) with respect to the concentration of microorganisms when the substrate concentration is large. Substrate utilization is second order (k' is the rate constant, equation [8]) with respect to substrate and organism concentrations when the substrate concentration is small (85, 87). The measurements of surfactant concentrations that are recorded in Tables 10, 22 - 24, 26, and 27 23 provide evidence that nonionic surfactant concentrations in surface waters, sediments, and groundwaters are usually well below 1 ppm. Therefore, the processes described above for secondary substrate utilization and co- metabolism generally apply for the transformation of surfactants. Many factors, not yet understood, affect the rates of biodegradation in primary and secondary substrate utilization, and co-metabolisrn. Enzymes, produced by microorganisms while meeting their metabolic needs, catalyze reactions, but a very complex set of rules has evolved to govern enzyme induction, repression, inhibition, and activation (85). Together with varying environmental conditions such as substrate mix, microbial population, organic content, nutrient availability, suspended solids, pH, temperature, acidity, and salinity, there are too many factors that can cause wide variations in biodegradation rates from one location and time to another to allow reliable prediction. In spite of the above limitations, some researchers have attempted to determine biodegradation rates using a more direct and simplified approach (75). The many complexities of biodegradation processes are ignored. The test compound is the sole source of carbon and energy (75, 88) and may be radiolabelled (14C) (75, 78, 86, 89-92). Compounds are tested in various environmental matrices. Over a range of concentration levels spanning up to several orders of magnitude, measurements of C02 or 14C02 evolution and C or 14C substrate removal have yielded curves that follow zero (78, 86, 90), first (75, 78, 86, 88-90, 92), or mixed-order ( 75, 86, 90) kinetics. At lower concentrations that are representative of environmental levels, C02 evolution and substrate removal are typically best described using first order kinetics, as depicted in Equations [9 -10] below. The rate constant is determined with a non-linear regression model. C02 or l“C02 evolution data and substrate (C or 24 14C) removal data is analyzed using a basic exponential product formation equation and an exponential decay equation, respectively (89-91). 0 Formation: P - 0 for t s c [9] P- P0(1 - e-th-Ci) for r> c where: P- C02 produced P0 - maximum amount of C02 produced (theoretical) t - time (days) c - lag time (days) It, - production rate constant (clay‘1 ). 0 Decay: C = Co for t S c [10] C= C0 (e'kzlt'cb for t> c where tand c are as defined above C = amount of substrate remaining in solution C0 as initial substrate concentration k2 = decay rate constant (day'l) First order equations similar to equation [9] have been found to describe the ultimate biodegradation of compounds in both soil and aquatic test systems (75 ). Equation [9] can describe ultimate biodegradation when environmental concentrations are low because net growth (production of biomass) occurs only at higher concentrations, after energy needs for maintenance have been fulfilled. In keeping with the previous discussion following equation [8] about the interrelationship of substrate utilization, substrate and organism concentrations, and the rate of biodegradation, the two rate constants, k, and 25 k2, are really pseudo first order rate constants, with k, = 19' X1, and k2 = 19' X2. That is, k; ' and 19' are second order rate constants with respect to substrate and organism concentrations (75, 91). When pseudo first order kinetics applies, another approach in safety assessment evaluations may consider the ratio of a compound's biodegradation half-life (BHL) and its residence time (RT) in a particular environmental compartment. The biodegradation half-life is the time it takes for half the substrate of interest to biodegrade. The residence time is the length of time the compound is in the particular environmental compartment of concern. The ratio - BHL/ RT - provides an estimate of the amount of biodegradation that occurs. If the ratio is equal to one, about half the material will biodegrade. As the ratio becomes very much smaller or greater than 1, then greater or lesser amounts of removal via biodegradation, respectively, will occur ( 76, 81, 93-95). Once the biodegradation rate constant has been estimated, the BHL can be obtained by solving equation [9] for t such that: t1); = 0.693/k, [11] Though calculation of the BHL/ RT ratio may provide a relative sense of the effect of removal by way of biodegradation, it cannot be the only criteria by which surfactant safety is evaluated. The value calculated for the biodegradation half life is only a first approximation, just as the rate constant calculated by fitting data to equations [9 and 10] is a first approximation (75), The residence time in a given aquatic or soil system can be estimated, but it can vary widely with the passage of time, from one location to another, based on flow velocity, settling velocity, suspended solids, and clay/organic content. 26 Other Factors Affecting Biodegradation Many other aspects must be considered in biodegradation studies. A compound's biodegradability or, conversely, its recalcitrance is a function of both the conditions of its environment and its molecular structure. Table 5 lists factors that affect the extent and rate of biodegradation of organic compounds in the laboratory and in nature. The absence of any one of these may prevent a compound from degrading. Table 5. Factors Affecting Extent and Rate of Biodegradation of Organic Compounds ‘ 1. Accessibility of substrate to biodegrading organisms 2. Presence of requisite enzymes - Enzyme must exist (assumes prior exposure to similar molecular structure) - Enzyme must be available to act on substrate - induced if it is not constitutive - not inactivated by adsorption or inhibition - Cell walls must be permeable to compound if enzyme acts intracellulariy 3. Availability of requisite substances for energy and growth (water, organics, nutrients, minerals, electron acceptors) 4. Presence of favorable environmental conditions (pH, salinity, alkalinity, temperature, osmotic pressure, toxicity) 5 Presence of molecular structure(s) that enzyme(s) can act on ‘ .Ref. (96-99) Physical-chemical properties also influence all reactions and processes that occur. Applied early in a new chemical's life, physical-chemical property constants provide a first determination of a chemical's behavior and fate by defining characteristics such as solubility, partitioning, and sorption (76). Constants that are derived in the laboratory, under highly structured and well controlled conditions, may not represent what occurs environmentally. Varying test conditions can yield widely different results. Constants for surfactant metabolites that have been determined in various studies are provided in Table 6. 27 Table 6. Nonionic Surfactant Physical-Chemical Properties Constant 5119119339991 up NPE, NPE; NPE3 NPE1-3C 09 Ref. stimulates Distribution coefficient 3 10,500 1,800 900 --‘ 500 --‘ (29) (um) Sorption constants ‘ 10,000 - 3,900 - 1,300 - --‘ --‘ --‘ (15) (L/kg) 26,000 1 1,000 6,900 .09 Kowz 4.2 --1 --1 "1 --‘ 3.7 (24) log Km 4.48 4.17 4.21 4.20 4.075 4.12 (100) (9 205° (2)2 10.12 10.15 10.18 10.11 10.10 log Km, 3.7 3.4 3.3 --‘ --‘ 3.0 (100) (G 20.5° C)2 10.04 (n-1) (n-1) 10.04 Aqueous solubility 5.43 3.02 3.38 3 5.88 --‘ --‘ (101) (mg/L 9 20.5°C) 10.17 10.07 10.12 10.13 Aqueous solubility 24.7 1 1.4 11.0 8 16.7 --‘ --‘ (101) (umollL G 20.5°C) 10.77 10.27 10.39 10.34 OP OPE; OPE; OPE3 Aqueous solubility 12.6 8.0 13.2 18.4 ( 101) (mg/L @ 20.5° C) 10.50 10.18 10.21 10.55 Aqueous solubility 61.2 32.0 44.9 54.4 (101) (umol/L Q 20.5°C) 12.43 10.72 10.71 11.63 Al 01 A E C10Ep CrzEp C1459 C15Ep ethoxylates p-3,5.8 p-3,5.8 p-3.5.8 p-S.8 Sorption constants 40-130 250- 2950- 4750- (sz)Z (ng) 1230 3550 6200 (102) log KSW 1.61 - 2.41 - 3.47- 3.68-3.79 (102) (@ 28° C)2 2.10 3.09 3.55 Sorption constants C15E9 (103) (K5102 (L/kg) 350- 21006 log sz 2.477 (76) ‘ Not available. 2 K.M - octanol-water partition coef.; Khw - hexane-water partition coef.; sz- sediment-water partition coef. 3 Ratio between concentration in activated sludge and 2° effluent. 4 As related to activated sludge samples. 5 Estimated; calculations provided in Appendix C; method in Ref. (35). 5 Range measured in 4 sediments. 7 Structure not identified; temperature not provided. 3 Ahel and Giger feel this result is low (experimental error); if adjusted to fit linear regression model the values would be 4.30 mg/L and 14.0 umol/L. 28 Experimentally obtained Kow values could not be found for the carboxylated alkylphenol ethoxylates. The Kow for NPE1 - 3C that is provided in Table 6 was estimated by adjusting the Kow for NPE1-3, 4.20, (100), based on a fragment constant that has been determined for a carboxylic acid group (35) . More details about the calculation are provided in Appendix C. Values for the aqueous solubility constants in Table 6 depict an anomaly that deserves comment. AP is considered to be the most lipophilic of the APE derivatives. This is corroborated by larger distribution coefficient and octanol—water partition constant (log Kow) values for NP than for the lower oligomers. However, the usual inversely proportional relationship between increasing distribution coefficient and log Kow and decreasing solubility does not apply. Instead, the water solubility values are considerably higher for NP and OP than for any of the lower oligomers. Ahel and Giger attribute this to the dominating role that the slightly acidic phenol group plays vs. the alcohol group in the ethoxylate chain (101). Another approach to solubility estimates considers the effects of pure rather than mixed chemicals on solubility. The solubility values in Table 6 were derived in pure water with pure chemicals. In order to more closely simulate environmental conditions, Ahel and Giger determine solubilities experimentally for a mixture of nonylphenolic compounds as found in two commercial surfactants, and arrive at very different results, Table 7. Their experimentally obtained solubility results compare closely with solubilities calculated using an expression presented by Stumm and Morgan (104, p. 734). Stumm and Morgan's expression calculates solubility for a mixture of chemicals, assuming that each individual component behaves ideally, according to: 29 Solubility ofmixture =- X181 «1» X282 + + XpSp = 2 XiSi, i -= l-p, [12] where X; is the mole fraction of component i, and Si is the pure water solubility of the pure component i. NP is not found in the commercial mixtures, but a look at the results for NPE1.4 demonstrates that an individual component's contribution to solubility is largely dependent on its mole fraction - the smaller or larger the mole fraction, the smaller or larger its respective contribution to solubility. Table 7. Comparison of Pure and Mixed Aqueous Solubilities of NPEM 1 Concentration (umol/ L) Pure2 Marlophen 833 lmbetin N/7A‘ Compound experimentals calculated experimental5 calculated NPE1 11.4 2.1 2.0 8.0 8.6 NPEZ 14 5.0 5.6 2.4 2.8 NPE3 16.7 4.7 4.7 0.5 0.8 NPE; 19.3 2.1 2.7 nd 5 nd 5 .2_XENPED Sol - - 13.9 15.0 10.9 12.2 1 Adapted from Ref. ( 101). 2 Values from Table 6. 3 Original mix mole fractions for NPE1, NPEZ, NPE3, and NPE4 - 0.174, 0.401 , 0.281, and 0.142, respectively. ‘ Original mix mole fractions for NPE1, NPEZ, and NPE3 - 0.75, 0.20, and 0.05, respectively. 5 Determined using generator column method. 5 Not detected. Sorption has a profound effect on the extent and rate of surfactant biodegradation. While biodegradation of a compound may ultimately lead to complete mineralization (conversion to C02 and water), sorption decreases bioavailability as a surfactant is transferred from the aqueous phase to a solid phase. Through sorption a compound is removed from the water column as it adsorbs to particulate material, then settles to become associated with sludge or sediment. When sorption is significant, its effects must be factored into safety assessments. A variety of approaches are found in the literature. Eadsforth and Moser, 1983, (1 05) compared chromatographic methods for determining partition coefficients for a wide variety of chlorinated and non-chlorinated 30 organic compounds. Abel and Giger (100) used HPLC to determine coefficients for low Oligomer APEs (Table 6), and proposed a method for estimating coefficients for high Oligomer APEs. Kiewiet et al. (102) used sorption data to write a quantitative structure-activity relationship (QSAR) equation that estimates a sediment-water partition coefficient based on alkyl and ethoxylate chain lengths. Cano and Dorn (103) compared sorption of an AB (C15E9) for various sediments. Brownawell et al. (106) characterized the mechanisms involved in surfactant adsorption according to molecular structure, solution composition, and sorbent properties. They determined adsorption isotherm constants for nonionic, anionic, and cationic surfactants on five sediment and soil types obtained from the EPA. Scow and Hutson (107) examined the relationship of diffusion and sorption to biodegradation kinetics while still others have considered the interactions between adsorption and desorption on partitioning in effecting removal from the water column (61, 108), Ideally, environmental studies should be executed in the field in the appropriate environmental compartments at realistic concentrations. This approach is impractical for satisfying all research needs related to surfactants because of the complexities of the interrelationships and the variability of conditions that are possible environmentally. However, recent improvements in analysis techniques provide more reliable results from environmental samples and have eliminated some of the uncertainty in the exposure part of the environmental risk assessment equation. After presenting proposed biodegradation pathways and a brief summary of conclusions drawn from pathway and kinetic research, findings from analyses of environmental samples with respect to surfactant behavior in appropriate environmental compartments will be reviewed. Blodegradatlon Pathways The literature reveals a variety of possible biodegradation pathways for AEs and APEs. This paper briefly discusses the pathways seen most commonly (Figures 5 - 9). Evidence of the mineralization of AEs or APEs by a single organism could not be found in the literature. It is more probable that the requisite enzymes for mineralizing nonionic surfactants are produced by a mixed consortium of microorganisms. Aerobic and anaerobic organisms that are capable of growing on nonionic surfactants or polyethylene glycols (PEGs) are listed in Table 8. If available, information about experimental conditions, original source of organisms, points of attack, and byproducts is provided. Candida maltosa is the only yeast in the list; it utilizes linear 4-1- nonylphenol as sole carbon and energy source by degrading the alkyl chain. Detailed coverage of reaction mechanisms and enzymology is beyond the scope of this paper. Reviews that offer additional perspectives can be found in Cox, (109), Cain, (2), and Swisher, (6). Evidence of nonionic surfactant biodegradation is available for a diverse set of compartments and conditions. Most surfactant biodegradation research has been conducted with mixed cultures enriched from activated or digested sludge (aerobic or anaerobic). Cultures from soils, sediments, and surface waters have been used to a lesser extent. Environmental compartments that have been simulated in the laboratory include: serni- continuous activated sludge (SCAS) system and simulated surface water from the system (88); continuous flow activated sludge (CAS) system (1 7, 74); sludge digester (110, 111); creek sludge (111); lake water (74); river water (75, 89, 90); estuary water (86, 90); marine water (112); sediment (106); and soil (92, 106). 31 32 Table 8. Organisms Capable of Growing on Nonionic Surfactants 0r PEG Remarks: Culture (Eure or fijxedfl Ref Conditions! Source/ Point of attack] ' Organism Comound Byproduct Pseudomonas P400 0121523 (Brij 35) P; aerobic; soil; (113) Pseudomonas sp. AE (CME9), NPE, P; aerobic; dyeing wastewater; (114) strain 14-1 Pseudomonas sp. NPEp P; aerobic; EO chain degraded; to 90% TOC - (115) authors speculate that NPE; accum'd, but provide no evidence; Pseudomonss sp. C5E3,C10E3 P; activ'd sludge; EO and alkyl chains (116) strain TR01 degraded; byproducts: short chain fatty acids; CieEzo P; activ'd sludge; partial EO chain degraded; OPE”, NPE” P; activ'd sludge; EO chain degraded via oxidative p’way (glycolate and glyoxalate assimilated); byproducts: NPE: and NPEZC; Nocardia R¢E5 (APEs) P; growing cells: aerobic; (9) R - H; CH3; resting cells: aerobic, anaerobic; R - n-C4 (lin); enrichment culture; . R _ sec-C4 (bf); EO chain only degraded; growth on lin C4 very slow; no growth on br C4; when R - H, H¢E1 accumulates; Cylindrocarpon H451 P; aerobic; enrichment culture; Pelobacter C15E20 (an 58), P; anaerobic; marine, Venice River sediment; (117) venetianus partial EO chain degraded; fate of alkyl chain not mentioned; K0835 Cro-12E7_5. M; anaerobic - fixed bed rctr; polluted creek (111, (~Pelobacter (2121523 (Brij 3s), sediment or anoxic digested sludge; 118) propionicus); APE (C1012¢59)3 K0835 and K0858 grow on E0 chain, K0858 studied in (111) M. hungatei 8: M. soehngenii on fatty acids; AE (~Acetobacten‘um byproducts: CH4, C02, traces of acetate and 59-); propionate; APE byproducts: CH4, acetate, ~Methanospirillum alkylphenol residues hungatei; ~Methanothn‘x soehngenii K0835, C12E23 (Brij 35), P; anoxic digested sludge; EO chain only (118) K0858 CreEzo (Brij 58), degraded; isolates did not reduce nitrate, (see above entry) p555 sulfate, sulfur, thiosulfate, or sulfite; byproducts: acetate, propionate (K0835 only), long chain fatty acids (from AE's); Candida maltosa 4-1-NP (lin) P; aerobic sludge from trtmt of textile plant (119) wastewaters; alkyl chain degraded; byproduct: 4-acetylphenol 33 Table 8. (cont'd) Methanobacten‘um, EG, DEG, 9E6 M; anaerobic digester sludge; (120) Desulfovibrio byproducts: acetate and methane; Methanosarcina sp. PEGs: 22EG, 450EG Desulfovibrio EG, DEG, 3156, P; anaerobic digester sludge; (121) desulfun‘cans, 4EG ether bonds cleaved by extracellular hydrolysis or hydrogenation; Bacteroides sp. PEGs: DEG to byproducts: acetate, ethanol, hydrogen; 450 EG; C1;E_z3 (Brij 35) Field studies have been conducted in SI'Ps (11, 13-16, 18, 23, 29, 39, 40, 49, 54, 55, 122-131), fresh surface waters (12, 14, 19-22, 40, 49, 54, 131-136), groundwater (62, 68), marine water (13 7, 138), drinking water (19-22, 40, 62, 132), and sludges, soils, and sediments (54, 62, 130, 133). The impact of environmental conditions on extent and rate of biodegradation has been studied by modifying the temperature (10, 1 7, 36, 41, 75, 89), substrate concentration (1 7), test inoculum (75, 88, 111), and dissolved 02 levels (75, 110, 111, 118). The latter three references provide results for studies completed under anaerobic conditions. Key points of these and other earlier works are presented in more detail in Swisher (6) and Talmage (27). Alcohol Ethoxylates Figure 5 depicts a general scheme for the biodegradation of alcohol ethoxylates. AEs can be attacked initially in one of three places (2, 6) by way of: I) rip-oxidation on the terminal methyl group of the hydrophobe end (77, 110, 139, 140); II) hydrolysis, directly or oxidatively, at the central ether bond joining the hydrophobic and hydrophilic moieties (74, 77, 86, 1 10, 139-141); 34 III) step-by-step elimination of the ethoxy (EO) units of the hydrophilic end (6, 113, 117, 124, 140. Further discussion about early pathway research is provided in Appendix D. CHa-(CH2)m-1 '0-(CH20H 20) pH 1 III \A.../\O/\/Ov ~-\/°\/fl II L 11 (aerobic) II l 1111 I. try-oxidation of II. central fission 111, attack of terminal hydrophobe, A. hydrolysis to monomer unit of followed by alcohol and PEG; ethoxy chain B-oxidation 8. oxidation of a-C A- direct hydrolysis; See detail, Figure 6. of hydrophobe; 3- modified hdeOUSlS ' dehydration-hydrolysis; C. dehydration-rehydration; D. dehydrogenation; See detail, Figure 8. C. oxidation of a-C of hydrophile; See detail, Figure 7. Figu'e 5. Major proposed biodegradation pathways for AEs. Molecular structure and environmental conditions, including the presence of oxygen and the nature of the organisms or enzymes present, determine whether attack is most likely to occur at point I, II or III (142) (Figure 5). It is generally agreed that the hydrophobic alkyl chain of AE degrades more quickly than the hydrophilic ethoxy chain, and that the rate of primary biodegradation is fastest when the hydrophobic moiety is primary and linear, which is usually the case for AEs used in detergents. Given an AB with a linear primary hydrophobe, attack is more likely to occur at points I or 11 than 111; yet no evidence could be found in the literature whether hydrolysis (I) or central fission (11) occurs more readily in the environment. Delayed degradation and much slower rates of degradation that occur with a secondary or branched hydrophobe (141) (deficiency of requisite enzymes) increase the likelihood that the ethoxy chain (III) will be attacked first. 35 W The reactions in Figure 6 are similar to those seen frequently in the literature for straight chain alkanes (143-149). The first orddation reaction activates the carbon of the terminal methyl group to allow insertion of an oxygen atom. Elemental oxygen, a hydroxylating NADH—dependent mixed function oxidase (monooxygenase), and a reduced cosubstrate (rubredoxin or cytochrome P450) must be present to initiate the first step of «ii-oxidation, which converts the terminal methyl group to an alcohol group. The next two oxidation steps are dehydrogenation reactions which convert the alcohol to an aldehyde, then to a carboxylic acid. The final step preparing the chain for 5- oxidation is esterification, which converts the carboxyl group to a thioester (- S-CoA) group. The alkyl chain is now prepared for B-oxidation, a sequence of four reactions that removes two carbon groups, the terminal S-CoA group and its a-carbon, from the side chain. B-oxidation repeats until there are one or two carbons remaining in the chain. In this example, the final intermediate shown is similar to monocarboxylated polyethylene glycol (PEG). 36 I CH3’(CH2)m-1 “04011203 20') pH 02-1-NADH4-H" \/\ ---/\o/\,0\/ VowOH \I CH210HHCHzirul-1 -0-(CH20H20-)pH NAD+ NAD“ H0\/\ "'/\0/\/°\/ VOWOH NADH+H+ NADH+H+ cozH-(CH21n1-1°0-(CH20H20-)pH S-CoA-i-ATP H020\:‘<\ "‘/\0/\/O\/ \/O\/\OH S. ! COzH-CH 2-0-(CH20H20)9H (Carboxylated PEG) . —'* —> HOZCAONOV \,0\/\0H p-oxidation repeats Figue 6. Proposed biodegradative rib-oxidation pathway for AEs. mm Figure 7 depicts three major alternative hydrolytic central fission pathways. The most direct pathway, designated A, is a simple hydrolytic reaction, which produces an alcohol and PEG. Hydrolysis also occurs in the remaining two pathways, but a hydroxyl group is added to the a-carbon of the alkyl or ethoxy end (designated 8 and C, respectively) before cleavage occurs. As shown, each step of the pathway leading to hydrolysis offers two alternative reactions. The hydroxyl group is added to the hydrophobic or hydrophilic moiety either by activation of the a-C and oxygen insertion similar to the first uroxidation step described above for the alkyl chain, or dehydrogenation at the 01-8 bond is followed by insertion of a water molecule. The hydroxylated intermediate is a hemiacetal which is unstable and may be easily hydrolyzed, forming an aldehyde and an alcohol. Alternatively, dehydrogenation may occur, converting the hydroxyl group to a carbonyl 37 group, thus forming an ester (1 0) which may be hydrolyzed. In this case the products are an acid and an alcohol. The specific products formed are dependent on which a-C is hydroxylated (alkyl or ethoxy chain) and whether hydrolysis of the hemiacetal or esterification followed by hydrolysis occurs. More information can be found in Swisher, ( 6). The products of two possible reaction sequences are shown at the bottom of Figure 7. Wuhan The final major pathway is depicted in Figure 8. It has been generalized to show that biodegradation of the ethoxy chain is thought to be similar for AEs, APEs, and PEGs. All pathways depicted in Figure 8 propose that the ethoxy units are cleaved by hydrolysis. As in Figure 7, the pathway designated A depicts direct hydrolysis. Non-oxidative or orddative modification of the ethoxy unit precedes hydrolysis in the remaining pathways. Pathways designated B and C are non-oxidative; the pathway designated D is oxidative. Direct hydrolysis Haines and Alexander (150) and Schink and Stieb (1 1 7) suggest that a direct hydrolytic cleavage reaction occurs in the biodegradation of PEG in both aerobic (150) and anaerobic (1 1 7) systems, though neither group explains the cleavage mechanism. Haines and Alexander experimented with a strain of Pseudomonas aeruginosa isolated from soil. By producing an enzyme that catalyzes depolymerization extracellularly, their organism can degrade mono-, di-, tri- tetra, and polyethylene glycols with up to 450 units (average molecular weight 20,000) . They identified diethylene glycol as a product of hydrolysis, and concluded that monoethylene glycol or possibly oligomers with 2 to 8 units of ethylene glycol may also be products. If hydrolysis is indeed the mechanism, 38 CH3'(CH2)m-1 -0-(CHzCH 20') pH V\ ---/~0’\,0\/ \OwOH II A. hydrolysis H20 CH3-(CH2)m.1-OH + B. oxid'n of a-C of hydrophobe by add'n of -OH rou l e 1 1C. oxid'n of (PC of hydrophile by add'n of -OH roup (alcohol) 1. dehydrogenation- I 2. hydroxylation I HO-(CHgCHzO)pH (polyethylene glycol) 82b C2D 81 a, 01a 1 Fig 8 , 52“ 81b ' , a. hydrolysis of C28 Clb b. esten'fication for PEG p way hemiacetal CHa-(CH2)m.1 -COOl-l fatty acid Products from 8 and C 1 - series of reactions: fi-oxidation B18, 82a: aldehyde + PEG 31b. 32b: fatty acrd + PEG Cia. C282 alcohol + PEG w/terminal C1 b, C2D: alcohol + aldehyde grp. carboxylated PEG Example of oxid'n of a-C of hydrophobe: CH3-(CH2)m. -O-(CH2CH20-)pH B1 2 H4- CH3—(CH2)m.3-CH=CH-O-(CH20H20-)pH B1 ”20 CH3-(CH 2)"..2 -CH(OH)-O-(CH 20H 20) pH 81a H20 CH3-(CH2)m.zCH(OH) + hemiacetal HO-(CH 20H 20) pH B1 3 H2 0 PEG CHa-(CHzi m-2'C(=O)H aldehyde See Fig. 8 fatty acid; B-oxid'n for PEG p'way Example of oxid'n of (rC of hydrophile: CH3-(CH 2)"..1 -O-(CH20H20-)pH C2 02; reduced cofactor CH3-(CH2)m.1-O-CH(OH)CH 20-(CH 20H 20) p.1 H 02b 2 H+ CH3 '(CHzim-l '0-C(=0)CH20-(CH201'120-) p-1H C2D H20 CH3-(CH2)m.1-OH 4» alcohol HOOC-CHZO-(CHZCHzO-)p.1H 1 carboxylated PEG fatty acid; 8- oxid'n See Fig. 8 for PEG p'way Figure 7. Proposed central fission pathways for AEs. 39 the hydrolase has never been characterized,. The organism was lost (1 1 7 ), and further work supporting their conclusions about hydrolytic cleavage by an extracellular enzyme could not be found. Schink and Stieb hypothesized two pathways, both intracellular, involving direct or indirect hydrolysis and ending in the production of equimolar amounts of ethanol and acetate from acetaldehyde via a disproportionation reaction. The direct hydrolysis pathway is shown in the anaerobic portion of Figure 8, pathway A. The strictly anaerobic organisms (Gra PEG 1, Gra PEG 2, and Ko Peg 2) used by Schink and Stieb consume dimers, oligomers (up to and including 8 units), and polymers with as many as 450 units of ethylene glycol. Schink and Stieb could not explain how the 20,000 molecular weight PEGs could permeate the cytoplasmic membrane. Monoethylene glycol (MEG) appeared to block the uptake mechanism and inhibited further degradation when any of the three strains were used alone. However MEG was consumed by a fourth anaerobic strain, Gra EC 12, when grown in coculture with Gra PEG 1, Gra PEG 2, or [(0 Peg 2. The strains were not able to reduce sulfate, sulfur, thiosulfate, or nitrate. The indirect hydrolysis pathway proposed by Schink and Stieb suggests that hydrolysis follows modification of the terminal ethoxy unit by a dehydration-rehydration reaction, discussed below and shown in Figure 8, pathway C. Ethylene glycol biodegrada tion A variety of possible pathways have been presented in the literature for the further transformation of ethylene glycol (EG). Anaerobically, or when the environment is stressed by minimal oxygen levels, 80 will undergo a rearrangement to 1,1-ethanediol (6, p. 678, Eq. 7.40 & 7.41, 118), Figure 8, pathway A. Schink and Stieb (11 7) present degradation steps for ethylene glycol under anaerobic conditions that include dehydration, enol-keto 4o tautomerization, and disproportionation of acetaldehyde to equimolar amounts of ethanol and acetic acid, shown in Figure 8 (following the rearrangement step in pathway A). Aerobically Wiegant and de Bont's, (151) pathway is similar, except that it does not include disproportionation. All of the acetaldehyde would be oxidized to acetic acid, then to acetyl-S-CoA, before entering the tricarboxylic acid cycle and the glyoxylate bypass. Steber and Wierich's pathway, not shown, proposes a series of four oxidation reactions, all dehydrogenations producing oxalate, which is further degraded to C02 and formate by a decarboxylation reaction (1 40). Steber and Wierich also consider the possibility of a pathway of a more circuitous nature, not shown, that leads to entry into one or a combination of other common metabolic pathways - the dicarboxylic acid cycle and glycerate pathway (140, 152-155), or Embden- Meyerhof-Parnas pathway (144). Modified hydrolysis Pearce and Heydeman, (156), isolated a strain of Acinetobacter, 88, that could effect cleavage, in the presence or absence of oxygen, in oligomers having between two and four units. Using diethylene glycol (DEG) as their starting compound, dehydration precedes hydrolytic cleavage (pathway B, Figure 8), producing ethylene glycol (EG) and acetaldehyde. Disproportionation as presented in the discussion above did not appear to occur. Not shown -- ethylene glycol accumulates; acetaldehyde is quickly converted to relatively large quantities of ethanol by an unknown reduction mechanism (aerobically and anaerobically) and smaller amounts of acetic acid (aerobically only). C2 units are further utilized in a variety of common metabolic pathways, as discussed above. 41 .81 as 8% .8... .o 5% 35o 2: to. £358 55283.. 8881 .o 2.51 1 u . 31.00.51. beanie . 1. ...éfiIOVbIOV 30.3 r m - a a 38820 0 <8 "cm 00.06 08.0820 .10.. 8.0. c.0188 + 8.101 89.8»...ooahx8mm + 01.0.”. 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SH IO/\/o\/ \/o\/\o/m 1201010101 3 12-0«1oa10.-o-.1o0-em .o 1a.-o~1ow10.-0-75310210 42 Dehydra tion -rehydra tion Another anaerobic scheme, hypothesized for linear alcohol ethoxylates (1 1 8) and analagous to a PEG biodegradation pathway proposed earlier (1 1 7) , involves dehydration and rehydration of the terminal ethoxy unit, forming a hemiacetal (pathway C, Figure 8). As with the central fission pathways, the hemiacetal could undergo hydrolysis or esterification before hydrolysis. Under this scheme, an alkyl-PEG or PEG, shortened by one ethoxy unit, and aldehyde or acetate is produced. Oxidative hydrolysis Oxidative hydrolysis begins with dehydrogenation that attacks either the terminal carbon (157, 158) or the 1-2 carbon bond (113) of the terminal ethoxy unit (pathway D, Figure 8). Payne and Todd, (157), were the first to identify a flavoprotein dehydrogenase that is active with a secondary alcohol ethoxylate possessing nine ethylene glycol units (secondary C1 1-15E9). Ferricyanide serves as electron acceptor. Payne and Todd were unable to identify products, and therefore made no conclusions about the reaction mechanism effecting cleavage. Nor did they suggest whether cleavage or dehydrogenation occurs first. However, they did determine that ethylene glycol is not metabolized by the dehydrogenase in extracts. Kawai et al., 1978, found a dehydrogenase linked with 2,6-dichloro-phenolindophenol (DCPIP) that oxidizes ethylene glycols with from two to 450 units. Attack of the aldehyde that is formed leads to a second dehydrogenation reaction that produces a carboxyl group on the terminal ethoxy unit. To lend additional support to this pathway, Watson and Jones, (159), and Steber and Wierich, (140) identified carboxylated ethoxy intermediates in their research. Watson and Jones were successful in degrading PEGs with up to 34 units (average molecular weight 1500); Steber worked with linear C13E7 and PEG 400 (average 43 molecular weight 400, nine units). Thélu et al., (113) identified a dehydrogenase that degrades PEGs up to PEG 400. Their dehydrogenase is DCPIP-linked, but initiates removal of one hydrogen from each of the 1- and 2- carbons, thus creating a double bond between the carbon groups in the terminal ethoxy unit. The addition of a water molecule across the bond introduces a hydroxyl group on the 2-carbon, thus forming a hemiacetal, which is either hydrolyzed directly or esterified prior to hydrolysis. Additional studies Research examining ether cleavage reactions has also been conducted for alkoxyalkanes (160), chlorophenoxy-acetates (161-166), chloromethyl- phenoxyacetates (163, 167), methoxybenzoates (168-1 70), and methoxyphenyl and methylene-phenyl compounds (1 71, 1 72) . Alkylphenol Polyethoxylates The major pathway in the biodegradation of the alkylphenol poly- ethoxylates, APEp, inferred from both lab (15, 23, 139, 1 73, 1 74) and field (15) experiments, is step by step elimination of the most terminal monomer unit of the hydrophilic ethoxy chain, Figure 9. As already discussed, the alkyl structure of the alkylphenol polyethoxylates is usually highly branched, with an average of eight (octylphenol polyethoxylate - OPE) or nine (nonylphenol polyethoxylate - NPE) carbon groups. Dirnethyl branches are not unusual, resulting in quaternary carbons that are not likely to be attacked readily by the same enzymes that effect conversion of an alkyl chain to a fatty acid (via «ll-oxidation) in preparation for B-oxidation (2). Unlike the alcohol ethoxylates, central fission does not occur, probably because such a reaction is blocked by the presence of the highly branched alkyl chain together with the aromatic ring (139). Swisher argues that steric hindrance alone (from the branched alkyl group) explains the resistance to central fission (6). n 00(01420-12-0), H aerobIWerobic M R O-(CHg-CHz-O-)p.1 CH2COOH R O-(CHz-CHz-O-h, H p -1-3(APE,,C) p -1-3(Apt,) + F1 0051120112011: H anaerobic wwt/sit p - 1-3 (APEp) anaerobic sit k R - branched alkyl, usually C8 or C9 F1 ‘O—OH wwt - wastewater treatment (AP) sit - sludge treatment Figure 9. Proposed biodegradation pathways for APEs. Adapted from (15). Mineralization of the APE structure, i.e. biotransformation to C02 and water, will occur only if enzymes are present to effect shortening of the ethoxy chain, elimination of the shortened ethoxy chain, ring cleavage, and finally oxidation of the alkyl branch. Evidence of the isolation of a single organism that produces enzymes capable of catalyzing all of these activities could not be found in the literature. Organisms that attack the ethoxy chain have been isolated, as depicted in Table 8. The mechanism for terminal ethoxy monomer elimination has never been definitely determined, but it is thought to be a hydrolytic (1 1 7, 1 40, 150) or oxidative-hydrolytic (1 5 6) reaction, similar to that already described for polyethylene glycols. Again, the reaction can occur aerobically (116, 140, 150, 156) or anaerobically (6, 11 7). Reported results vary, but generally removal of the ethoxy monomers continues down to one, two, or three ethoxy units (15, 23, 1 74). Under aerobic conditions carboxylation may occur either on a terminal carbon group on the alkyl end (provided that the hydrophobe is not heavily branched), (139, 1 74), or on the 4S terminal ethoxy group (15, 23, 116, 126, 127, 1 75). APECs with over 3 EO units have been detected in environmental samples, raising speculation about when carboxylation occurs (1 8, 20, 23, 132, 1 76). The lower ethoxamers, whether carboxylated or not, tend to accumulate prior to being biotransformed during the next stage of biodegradation, conversion to an AP or carboxylated AP. The alkylphenol, if formed, is also resistant to biodegradation. Giger et al. found that NP accumulation is greatest in anaerobically treated sludge (13). Research results providing evidence for a mechanism that transforms branched alkylphenols could not be found. Steric hindrance most likely prevents attack of the alkyl moiety. If biodegradation proceeds further, it is probably by way of oxidation of the aromatic structure, which prepares the ring for cleavage by transforming the AP to an alkylcatechol (144). The branched substituent may be degraded by a combination of steps involving disruption of the quaternary structure (1 77, 1 78) and B- and (ll-oxidation steps (147). ' Recently Corti et al., 1995, (1 19), isolated a strain of the yeast Candida maltosa, LMAR 1, from aerobic sludge samples, that transforms a linear alkylphenol, 4-( 1-nonyl) phenol (Table 8). The predominant metabolite is 4- acetylphenol. Corti et al. conclude that a combination of B-oxidation and decarboxylation reactions occur to reduce the length of the alkyl chain. At some point, further degradation must proceed via aromatic cleavage. Though other metabolites are present, but not identified, it is unclear whether or at what stage during biodegradation the aromatic ring is cleaved. Pathways, Extent and Kinetics of Biodegradation - Generalizations The material covered thus far is summarized below. Figure 10 is a diagram of the relative ease (rate and extent) of biodegrading several AE and APE surfactants. It is a pictorial representation of the narrative that immediately follows. The comparisons are relative, not absolute. Note that the surfactants are grouped into three clusters: 1) linear ABS and ABS with a single methyl branch; 2) linear secondary AEs, very high Oligomer linear PAEs, linear APEs, and branched PAEs; 3) quaternary branched ABS and APEs; branched APEs. The position of the branched PAEs in the second cluster is based on the assumption that they are highly branched. A minor amount of branching could easily place them at the beginning of the second cluster, before the linear secondary AEs. Legend: m - 8 - 18 C E . lin - linear m 3-20 . sec - secondary Cm(C1)EB- 20 :C m(C1)E1 2 br - branched qu - quaternary C"'E"‘°'$¢1<=-Clnfil 20 l I ' lin/511353-20 l CmE>20-30 linAPE1 2 : bl(311153-20 : l | : QUbr‘Cm 1-20' Extent of branching brCmE12 brAPészo I determines rate of brARE biodegradation. 1'2 quTAPE 1-20 Least resistant ------------------------------------------------- ~> Most resistant Figure 10. Relative ease of biodegrading AEs and APEs. Adapted from Ref. (46, 179). 47 The following conclusions can be drawn: 1) The structure of the hydrophobe (linear, branched, phenolic) determines the rate and extent of biodegradation (10, 1 7, 46, 141, 1 73, 180). a) linear and singularly branched hydrophobes: i) linear AE surfactant molecules undergo rapid and extensive primary biodegradation with removal of the alkyl chain (1 0, 1 7, 74, 140, 141); the length of the alkyl chain generally has little effect on the rate and extent of biodegradation of linear primary AB (10, 89, 90, 141, 179); ii) a single methyl branch on an alkyl chain does not significantly affect the rate or extent of biodegradation (10, 1 7, 1 79); iii) a primary linear AE will biodegrade more rapidly and to a greater extent than a linear secondary AB (10, 139, 141); b) the presence of a phenyl group in the hydrophobe slows the rate and decreases the extent of biodegradation significantly (139, 141, 1 73, 1 79); a hydrophobic structure with branching and a phenyl group degrades very slowly and incompletely (10, 139, 1 73); c) as hydrophobic branching increases, the rate and extent of biodegradation continues to decrease (1 7, 141, 1 74, 1 79); a hydrophobe with a phenyl group and quaternary branching is most resistant (141, 1 74). 2) Degradation of hydrophilic moieties of ethoxylate (EO) surfactants is slow to begin and proceeds slowly once begun (10, 74, 140). a) EO chains with < 20 units tend to biodegrade more rapidly and extensively; biodegradation is slower and less extensive when the b) 43 chains are more than 20 or 30 E0 units long (10, 1 79); however Iarson did find comparable biodegradation rates for long (22 E0 units) and short (3 and 9 EO units) AE homologs (89); the E0 chain associated with a linear hydrophobe will biodegrade significantly more rapidly and extensively than the E0 chain associated with a branched, phenolic hydrophobe( 10); APEs with very short EO chains (< 3 EO groups) are most hydrophobic, most likely to partition to organic matter, and therefore least likely to biodegrade. Environmental Analysis and Safety Assessment The development of specific and sensitive analysis techniques, for the detection and quantitation of nonionic surfactants in environmental samples at very low concentrations, has been very difficult (181). Analysis of the polyethoxylate nonionic surfactants is more complex than analysis of anionic and cationic surfactants because: 1) the number of possible molecular structures is greater; 2) analytical response of intermediates cannot be easily differentiated from the parent surfactant; and 3) interference from other substances is frequently a problem. (6, p. 140) Within the last several years, techniques have been developed that can be used to determine the compositions of both the alkyl and ethoxylate moieties of alcohol ethoxylates at very low concentrations in environmental samples (41, 181, 182). Recent reviews summarize the extraction/enrichment, clean- up, separation (chromatographic), and detection methods that are currently available to complete specific and sensitive analyses of: 1) AEs in aqueous and solid environmental matrices ( 38); and 2) AE and APE homolog, isomer, and Oligomer distributions in environmental aqueous samples (182). The reader is referred to these articles if more information is desired regarding current nonionic surfactant chromatographic analysis techniques. Concerns about environmental persistence and accumulation of APE metabolites to toxic levels, particularly alkylphenols and alkylphenol mono- and diethoxylates, compelled much earlier development of analysis techniques for detecting APE and its metabolites in environmental samples. Nonylphenol polyethoxylates have received the greatest amount of attention because as the 49 50 most widely used of the APEs they are most likely to enter the environment as a result of their use in detergent and industrial products (27, 39) Walter Giger, working with other researchers, has been studying the behavior and environmental fate of alkylphenol ethoxylates since the early 1980's (1 1-13, 55, 126, 127, 129, 130, 183). They observe that the APEs and their intermediates can be categorized into four major groups according to their behavior - the parent alkylphenol polyethoxylates (APE3-20), alkylphenol mono- and diethoxylates (APE1& 2), alkylphenols (AP), and alkylphenoxy carboxylic acids (APE1&2C) (15). Hydrophobicity, sorption and toxicity increase as the ethoxy units are eliminated from the hydrophilic end, down to two or fewer ethoxy units. Sorption can become more important than biodegradation in determining the fate of APE metabolites, particularly if the concentration of suspended solids is high. Aided by the development of more sensitive analysis techniques, which allow for the detection and quantitation of organic compounds in a variety of environmental matrices at ever lower concentrations (Ppb. ppt, ppq) (40, 49, 131, 1 76, 181, 184), more attention is being devoted now than ever before to the removal processes that determine environmental behavior and fate. Results of field studies support the findings of laboratory studies related to surfactant pathways, biodegradability, and kinetics (18, 40, 75, 7 7, 78, 185). Findings from field studies are reviewed next, with respect to the appropriate environmental compartments of concern. 51 Alcohol Ethoxylates Alcohol ethoxylates easily pass stringent laboratory screening tests for ready biodegradability, and therefore are considered to be environmentally safe. Due to their reputation and the relatively recent development of analysis techniques that can be used to correctly identify a variety of AE homologs and oligomers at the very low concentrations present in environmental samples, AE environmental studies have been less extensive than APE studies. AEs are increasingly used to replace alkylphenol polyethoxylates in domestic, institutional, and industrial products as concern about the environmental effects of APEs becomes greater. As the use of AEs increases, questions are being raised about their environmental safety if used in much larger quantities on a widespread basis. Highlights from the available AE environmental studies follow. S w di ' i n 5 ms Matthijs et al. (78) used results from lab and field studies to demonstrate the importance of the sewer in removing a representative sampling of surfactants. They completed sewage die-away studies in the laboratory with AEs, alcohol ethoxy sulfates (AESs), and diethanolester dimethyl ammonium chloride (DEEDMAC) - a cationic surfactant. Half-lives and rate constants were estimated by nonlinear regression analysis assuming first order kinetics (Table 9A). Supporting evidence for results from the laboratory studies is provided by comparing actual measured environmental concentrations of ABS, glucose amide (GA - a nonionic surfactant), and boron to predicted concenuations based on consumption, water usage, and population data for the Netherlands. Measured concentrations for ABS and GA are each 47% lower than the predicted concentrations (Table 9B). Confirmation of this procedure lies in the fact that the concentration measured for boron, as a conservative 52 tracer, is very close to the predicted level (within 1%). AEs would be eliminated in the sewer at least to the same extent as AESs, given their lower estimated half-life and the fact that lab studies tend to be a conservative estimate of what is likely to happen environmentally (1 85-1 8 7). The half-life for complete biodegradation of ABS in activated sludge plants is estimated to be less than one hour (78). Table 9. Environmental Analyses in Sewer Distribution Systems1 Component Observed Estimated A) m Kinetics 1/2-life (hr)7- Lab Test Analysis AE (C14E3) zero order .. 3 sewage die-away TLC/RAD“ AES (C14E3S03) first order .0 4 sewage die-away TLC/RAD“ DEEDMAC mixed 6-10 sewage die-away TLC/RAD“ Predicted Measured Concentration Concentration 3) ELEM (pg/L) 3 (pg/L) Removal (36) Analysis AES 3400 1800 47 LC/MS5 GA 426 227 47 FAB/MSG Boron 893 900 0 OES7 1 Ref. (78), Netherlands; statistics not provided. 2 First order kinetics assumed. 3 Based on consumption data, population count, and water usage. 4 Thin layer chromatography with radioscanner. 5 Liquid chromatography with ion spray mass spectrometry. 6 Flow injection fast atom bombardment with mass spectrometry. 7 Inductively coupled plasma optical emission spectroscopy. The above approach does not consider removal that occurs in the sewer lines due to sorption. A more valid approach to ascertain that biodegradation occurs in sewer distribution systems would be to determine the change in total concentration by measuring concentration levels of dissolved and particulate fractions for the organic compounds of interest at both input and output points in the sewer distribution system. 53 W The following generalizations are highlights from three research studies that were conducted recently in four, three, and seven municipal, mechanical-biological sewage treatment plants in the United States (49), Italy (40), and the Netherlands (125), respectively. The U. 5. study (49) investigates removals in two activated sludge (A/S) and two trickling filter (TF) plants. The biological means of removal is nor identified for the study performed in Italy (40). All seven plants included in the study for the Netherlands are activated sludge plants (125). The results for one of the TF plants are excluded from the general summary statements in this paper because a major rain event that occurred before and during the sampling period affected measurements adversely. The data is, however, included in the tables and figures below. Environmental levels cannot be directly compared because test conditions, sampling methodology, analyte measurements and analysis techniques are different (27). Primary treatment Alcohol ethoxylate concentrations entering municipal sewage treatment plants vary from less than one to less than 7 mg/L (Table 10, Figure 11) (40, 49, 125). Average removal during primary treatment can vary from 6 (49) to 28% (49, 125). The 6% reading was obtained for the Oskaloosa, Iowa TF plant, the U. S. plant that did not experience the rain event mentioned above. Average removal during primary treatment is highest in the seven Dutch plants, averaging about 28% of the total influent load (125). AB concentrations were measured in the primary effluents for only five of the seven plants. Removals for the five plants range between 10 and 41%. Extent of removal is dependent largely on the settleable solids in the primary clarifier (49, 125), although with a two hour average detention time (Table 3) biodegradation also S4 contributes to removal. During primary treatment, the AE concentrations are reduced to between less than 1 and about 4 mg/I. (Table 10, Figure 11) (49, 125). Removal efficiencies during primary treatment are not provided by Crescenzi et al., perhaps because primary treatment is not provided in the Italian plants. Table 10. Environmental Analyses in Sewage Treatment Plants and Receiving Waters (09/ L) Recv'g Raw infl 1' effl 2" effl water low/high low/high low/high low/high Tap Comp (avgtsd) (avg1sd1 (avg1sd1 (avg1sd) water Analysis Ref. lLSA‘ C1z-15Ep 3210/ 2670/ 11/71 18/37 - - GC/MSD (49) (A/S) 3670 2820 w/HBr (34401 (27451 (41142) (28113) deriv'n3 325) 106) €12-15Ep 680‘/ 9504/ 504/114 30/34 - - GC/MSD (49) (TF) 2670 2500 w/HBr (16751 (17251 (82145) (3213) deriv'n8 1407) 1096) Italy? €17,138, 217.8/ - - 0.03/ 2.52 6 0.024 5 LC-ES-MS9 (40) 11515 12.865 (508.7 (4.0149) 1312) Netherlands3 C12-15Ep 1190/ 1020/ <3.0/ - - - - RP-HPLC (125) * 5500 34407 1 1.55 w/pic (2978 (1949 (<4.5 deriv'n; 11236) 1646) 12.0) UV det'n10 C1248Ep 1480/ 1240/ <4.1/ - - - - RP-HPLC (125) 6350 39307 21.75 w/pic (341 S (2290 (<7.5 deriv'n; 11365) 1743) 14.0) UV det'n10 1 lnfl & effl: 24 hr, 3 day flow-weighted composite samples from 2 activated sludge (A/S) and 2 trickling filter (TF) plants (late summer, early fall); rec'g waters: grab samples. 7- lnfl & effl: 24 hr composite samples, 2 or 3 days, from 3 mech'l-biol'l plants; rec'g waters: one grab sample. 3 24 hr composite samples, 3 consecutive days, from 7 A/S plants; measured aqueous 81 solid fractions. 4 Low TF readings occurred during rain event. 5 Different dates and plants for high 8: low readings. 6 One sample. 7 Five plants sampled. 3 Gas chromatography with mass selection detection; H8r derivatization. 9 Liquid chromatography coupled to mass spectrometry via an electrospray interface. 1° Reversed-phase high pressure liquid chromatography; phenyl isocyanate derivatization; UV detection. SS 3500 432:1. Q 3000 .. g. 2500 A 2000 223;; 1675 1725 1500 “2 .. .25; 3 Concentration ( C‘s" 1‘1" . 3:33: : : ::::: USA-A/S USA TF ltaly d (:1 2455;, C12-15Ep C12-18Ep C12-15Ep C12-18Ep Raw infl E31 1° effl D 2° effl Figure 11. Concentrations of AEs in raw influents, and primary and secondary effluents. Values from Table 10. Sorption is a large factor in removal from the water phase, as more than 50% of ABS can become associated with the solid phase (Table 11, Figure 12) (125). Recent investigations of ABS attempt to determine whether the length of the alkyl or ethoxy chain affects sorption. Fendinger et al. (49) conclude that an AB with a long alkyl chain is no more likely to be removed by sorption than an AF with a short alkyl chain. However their research is limited to AE molecules having between 12 and 15 carbon groups. Crescenzi et al. and Kiewiet et al. address behavior of AE molecules with up to 18 carbon groups. They observe that the longer alkyl chains, C16 and C13, are more likely to be associated with the solid phase and thus exist in the aqueous phases at lower concentrations (40, 125). Table 11 summarizes data (125) that demonstrates the importance of sorption as a removal mechanism during primary treatment. A greater fraction of the AEs is found in the aqueous phase of the primary effluent after solid settling. 56 Table 11. Liquid/Solid Phase Fractions of Raw and Settled influent 1 Raw influent Settled influent Range: liq/sol 2 Range: liq/sol 2 Component gAvg1sd: liq/sol} 3 (Avg1sd: Iig/sol) 4 C1z-C15 0.28-0.37/0.72-0.63 0.44-0.61/0.56-0.39 (0.3310.03/0.6710.03) (0.5210.07/0.4810.07) C1z-C18 0.26-0.35/0.74-0.65 0.38-0.57/0.62-0.43 (0.3010.03/0.7010.03) (0.4710.08/0.5310.08) 1 Ref (125). 7- Different plants for settled vs. raw influent. 3 n-6. ‘1 n-4. Solid D Liquid Raw infl Raw infl Settld Settld C1245 C1248 infl infl Clz-ls C12-18 Figure 12. A greater proportion of AEs is in the aqueous phase following 1° treatment. Values from Table 1 1. Secondary treatment Secondary treatment reduces AE concentrations to between less than 1 and 1 14 [lg/L (Table 10) (40, 49, 125). Biodegradation is the major means of removal during secondary treatment. Average removal efficiencies during secondary treatment range from 95 (49) to over 99% (40, 125). Chromatograms and related mass spectra of the oligomeric distribution of ABS in raw and untreated water suggest to Crescenzi et al. that central fission and oxidation of the alkyl chain, not ethoxy unit elimination, are the most likely biodegradation pathways. Radiolabeled compounds were not used in any of the 57 three studies to determine C02 production or biodegradation pathways. Fendinger et al. used 14C labeled compounds, C13E3 and C13E9, to evaluate the efficiency of the extraction and extract cleanup steps in their study (49). The removal efficiency of linear primary alcohol ethoxylates is not affected by lower temperatures in laboratory (1 7, 36) and field (188) studies. Stiff reports on the results of a study using the porous pot technique with C12- 15E9 and C9-1 1E3 (36). The AEs are removed efficiently under both winter and summer conditions, which were simulated by running tests at three temperatures - 8, 12, and 15°C. Kravetz et al. decreased temperatures from 25 to 15 to 8°C while testing linear €12-15E7, branched C13E7 and NPE9 in a continuous activated sludge flow-through system (1 7). Effluent CTAS values are constant at all temperatures when €12-15E7 (lin) is biodegraded. On the other hand, CTAS values increase significantly for C1387 (br) or NPE9 (br) when the temperature is dropped to 8°C. Turner et al. report on results of a field trial in biological filters (188). Influent and effluent temperatures were below 12.2°C throughout their study. Within the filters, temperatures between S and 10°C were measured. Yet the removal efficiency is between 96 and 98% for C14-15137 and C14-151311. Surface waters Thus far AE levels in natural waters have been well below toxic thresholds (27, 142, 189), and relatively little effort has been devoted to their study. AE concentrations measured in environmental samples of surface waters are provided in Table 10 (40, 49). Not surprising, lower concentrations are observed in receiving waters that have higher stream flows (higher dilution factors) (49). Environmental levels cannot be directly compared because test conditions, sampling methodology, analyte measurements and analysis techniques vary (27). Other researchers have also detected alcohol 58 ethoxylates in river water and drinking water (19, 20, 22), although quantitative information is not available. To determine whether or not alcohol ethoxylates biodegrade at trace concentrations after release to the environment, Larson and Games (90) and Vashon and Schwab (86) conducted laboratory tests with environmental levels of ABS in fresh water and estuarine water, respectively. Larson and Games (90) observed the biodegradation of 1‘1C-labeled compounds in Ohio River water. The compounds, C1 2139 and C16E3, were each radiolabeled either at the a-carbon of the alkyl chain or uniformly labeled in the E0 chain, resulting in four separate groups. From measurements of the rate of 14C02 production, they determined biodegradation rates for each of the four materials separately. From 14C activity converted to 1‘1COz, remaining in solution, and incorporated into biomass, they concluded that biodegradation of all four AE groups follows first order kinetics, occurs rapidly, and, is essentially complete. The rate and extent of biodegradation of the alkyl and EC chains for C16E3 are about equal. The alkyl and E0 chains of C12E9 biodegrade to the same extent, but the rate of 14C02 production with the C12E9 EO chain labeled is half the rate of 1‘1C02 production with the C12E9 alkyl chain labeled. The biodegradation half-life based on the mean value of the biodegradation rate constant of all four groups is approximately 35 hours. Vashon and Schwab (86) conducted a similar experiment with 14C- labeled AE compounds at trace concentrations in Escambia Bay water. They used C12E9 labeled uniformly in the 80 chain and C16E3 labeled at the a-carbon of the alkyl chain. Again biodegradation is rapid and extensive; first order kinetics apply at environmentally relevant concentrations. Central fission is indicated as the primary pathway. Biodegradation rates are lower in salt water 59 than in fresh water. Calculated half-lives equal 2.3 and 5.8 days for the respective C16E3 alkyl carbon and C12E9 ethoxylate carbon compounds. Sorption also plays a role in the fate of ABS in surface waters. Kiewiet et al. (102) determined sediment-water partition coefficients for a variety of ABS (€10-15E3-g) and found that sorption increases with increasing alkyl and ethoxylate chain length (Table 6). However the increase is greater with additionof carbon groups to the alkyl chain than with the addition of ethoxy units. Though Cano and Dorn (103) agree with Kiewiet et al. that sorption is a hydrophobic mechanism, they found that sorption decreases with increasing ethoxylate chain length. This is reasonable since more oligomers would increase the polarity, thus increasing the solubility. They examined the sorption behavior of C15E9 to four different sediment samples (Table 6). They did not compare AEs that vary in size hydrophobically. More studies are required to further evaluate effect of molecular structure and environmental conditions on biodegradation and sorption. Groundwater Evidence of a study conducted to determine the behavior and fate of alcohol ethoxylates in groundwater could not be found, no doubt due to the reputation that AEs enjoy as readily biodegradable compounds in STPs and surface waters. Research conducted by Knaebel et al. in 1990 (92) provides evidence that AEs biodegrade readily in soils that have not been previously exposed to anthropogenic surfactants, as well as in sludge amended soils. Knaebel et al. applied low concentrations of a 70%-30% mixture of C12E9 and C12E3 AE surfactants to 11 different soil types gathered from various U. 5. locations. They used surfactants that were uniformly labeled in the ethoxylate chain to track 14C02 production during mineralization. Measured C02 production was between 30 and 69%. The application of four wetting/drying 60 cycles increased the rate of mineralization. Similar behavior can be assumed in the event that AEs are introduced to groundwater. £11m Removal of alcohol ethoxylates begins in sewer distribution systems. AEs are removed efficiently during primary and secondary treatment in sewage treatment plants. Average total removals (primary and secondary treatment) are high, ranging from 96% (TF plant in the U. S. (49)) to almost 100% (activated sludge plants in the Netherlands (125)). Measured removals confirm quantitatively (40, 49) the high removals predicted in laboratory biodegradation testing and bench scale activated sludge system tests (1 7, 74). Central fission and (ll-oxidation of the hydrophobe are more likely to occur than step-by-step elimination of the ethoxy units (40). Lower operating temperatures do not decrease the efficiency of removal. Alcohol ethoxylate concentrations in surface waters in the U. S. and Italy are less than 40 113/ L. Laboratory studies indicate that alcohol ethoxylates biodegrade rapidly and extensively in both fresh and estuarine waters at these concentrations. The presence of alcohol ethoxylates in groundwater is not a concern environmentally. More studies are required to further evaluate the effect of AE molecular structure and environmental conditions on the interrelationships between biodegradation, sorption, and AE removal in STPs and surface waters. 61 Alkylphenol Polyethoxylates Worldwide agreement does not exist regarding the extent to which APE metabolites accumulate (41, 54, 116, 124, 127), thereby effecting ecotoxico- logical damage (29, 54, 62, 131, 133). Giger and co-workers have concluded that the accumulation of NPE metabolites, particularly NP, creates an environ- mental risk. Kubeck (131) and Naylor (5 4) have concluded that NPE meta- bolites do not accumulate and are not an environmental risk. Conclusions and an analysis of results of studies from the two viewpoints are below. Highlights of other investigations by other researchers, in other geographical areas, are also provided. In terms of the four major categories identified previously, researchers generally agree that the parent surfactant (APE3-20) is removed, while the metabolites (APE1&2, AP, APE1&2C) persist to varying degrees during treat- ment. The investigations of Ahel, Giger, and co-workers into the behavior and fate of these four groups allowed them to determine: 1) varying behaviors dependent on wastewater treatment conditions (15, 16); 2) mass and molar fluxes (29, 39, 128); 3) solubility constants ( 101) and partitioning coefficients (100). Most recently Ahel, Giger, and co-workers published three articles which are comprehensive summaries of their accumulated data and their conclusions about the behavior and fate of APEs and refractory metabolic products (29, 62, 133). The three phases of their study examine fate and behavior of nonylphenol polyethoxylates in sewage distribution systems and treatment plants (29), river water (133), and groundwater (62). Most of the generalizations that follow are taken from their three part comprehensive study. Ahel et al. demonstrated that biodegradation of NPEs begins in the sewer with data collected from the Uster sewage treatment plant (STP) (29, Figure 3). They compared distribution curves for raw influent, primary and secondary effluent, and a commercial NPEg mixture (Marlophen 810) used most commonly in laundry detergents (Figure 1 3 ) . They measured molar concentrations for each oligomeric configuration, NPF0-13. The distribution of ethoxy (EO) groups per molecule of nonylphenol for Marlophen 8 10 resembles a normal probability distribution, with a peak at 9 EO groups. Less than 1% by weight of the Marlophen 810 molecules carry one or two ethoxy groups (127). NP and NPEC are not present in the commercial mix (29). The raw influent, however, is distributed birnodally such that two peaks are obvious, at NPEl (highest peak) and NPEg. NP, NPE1, and NPEz are considerably higher in the raw influent than in the original surfactant mix: 260 vs. 0 umol/m3, 530 vs. 20 umol/m3, and 180 vs. 20 umol/ m3 for NP, NPE1 and NPEz, respectively. NPE3 is slightly higher at 55 vs. 40 umol/m3. The distribution of the NPE4-20 oligomers is similar to that for the commercial mix, but is shifted down and to the left. While providing evidence that biodegradation does begin in the sewer prior to entry into the water treatment plant, this also corroborates the proposed biodegradation pathway for the APEs, that is, step-by-step elimination of the ethoxy units according to: NPE] ,2C NPE3-2o —' NPELZ <: 1 NP e neutreetee eeueweter A Welly treated wastewater I , O biologically treated weflfletfl X emote! uneo mettle v §‘ \ \ K Concentration lirmoll m3) \K‘ ~ \ I / / O x' ' 3 \~.-. / . - - \ b . ‘\\‘\‘ O I \‘x‘Qx‘ . lzltstrrsou on it number of E0 groups per molecule Figure 13. Distribution of individual NPEfl oligomers in raw sewage, primary effluent and secondary effluent of the sewage treatment plant Uster and a commercial mixture Marlophen 810. Reprinted from (29, Figure 3), with permission from M. Ahel. A more accurate means to determine whether or not biodegradation occurs in the sewer is to compare measurements of the concentrations of nonylphenolic compounds in raw sewage samples at sewer input ports with concentrations at sewer outlets. Measurements of this type could not be found in the literature. W The bulk of data analyzed in the first phase of Ahel's 1994 report was collected during the 1980's from 1 1 plants located along a 35 km stretch of the Glatt River in Switzerland. At times Ahel and co-workers chose data from an individual plant or from two plants so as to make comparisons. Two-hour composite, 24-hour composite, and grab samples were measured. Previously published data is included in the tables and figures below for illustrative purposes; the nature and origin of the data is provided as appropriate. 64 The values provided in Tables 12 and 13 are rough estimates that have been derived based on information taken from several sources. Table 12 provides molar flows (mol/ day) for the Ziirich-Glatt STP. The values for the primary and secondary effluents are based on measurements from 24-hour composite samples ( 29, Table 4). The total raw influent value is estimated based on Ahel's assumption (29, Figure 8A), that 20% of total influent nonylphenolic compounds (NPc) settles during primary treatment and is routed to the anaerobic digester in the primary sludge. The raw influent values for the four major categories are estimated based on data provided and statements made in the text of articles (15, text, 29, text and Figure 3). The values for digested sludge are estimated based on grab sample measurements of total nonylphenolic derived compounds and nonylphenols in Ziirich-Glatt and Uster (29, Table 8). Figure 14 is a bar chart of Table 12 values. Table 12. NPE Molar Flows in lunch-Glatt ST P (moi/day (% of Total)) Raw infl 1 1° effl 2 2° effl Z Dig'd sldg 3 Category moi/d_ay(%) mol/day(%) mol/day(%) mollday(%) NPE3-20 83. (63.) 76.1 (55.9) 6.3 (4.6)5 tr6 NPELZ 40. (28.) 20.2 (14.9) 4.0 (2.9) 2.0 (1.5) NP 10 (7.) 6.3 (4.6) 0.9 (0.7) 29.0 (21.5) NPELZC 3. (2.)4 6.3 (4.6)‘1 19.8 (14.8) tr‘5 Tot'l NPc 136. (100.) 108.9 (80.0) 31.0 (23.0)7 31.0 (23.0)8 1 Ref. (29, Figure 3): Values estimated. 2 Ref. (29, Table 4): Table 4 value * 45,000 m3/day. 3 Ref. (29, Table 8 and Figure 8): Values estimated. 4 Ref. (15, Figure 12); slightly higher levels of NPELZC measured in Uster 1° effluent than raw influent. 5 NPE9_20 not present; NPE3- 8 present in trace amounts. 6 Only trace levels present. 7 Discharged to surface waters. 3 Usually land applied, or land filled. Molar Flows in Zurich-Glatt ST P 140 v 120 405-331 a 100 5 W MNPEIJC g so I I NP 5 60 i‘: NPELZ g ‘10 1211313320 20 Raw 1° 2° Dig'd infl't effl't effl't sludge figure 14. NPE molar flows in the lunch-Glatt STP. Values from Table 12. The reader is cautioned that measurements from composite and grab samples do not correlate. For example, composite sample measurements indicate that the overall elimination of NPc during secondary treatment in the Zi‘lrich-Glatt STP is 71% (29, Table 4); grab sample measurements indicate that the overall elimination of NPc during both primary and secondary treatment is only 58% (29, Table 8). The estimates of distribution ranges provided in Table 13 are an attempt to account for this. Also, Ahel et a1. (29) estimate that the environment ultimately receives 60-65% or more of all NPc (molar-based) that are introduced to sewage treatment plants. Between 30 and 40% of this (20—25% of total NPc) is released in the form of digested sludge; the remainder (35-45% of total NPc) is discharged in secondary effluents to surface waters. Table 13 values predict a less severe environmental impact than those estimated by Ahel et al. The values in columns 2-4 of Table 13 consider the elimination data of 24-hour composite samples of 1 1 Glatt River treatment 66 plants (29, Table 4); the digested sludge values in column 5 consider results from grab samples (29, Table 8). Ahel et al. found that the overall NPc elimination efficiencies during secondary treatment in the 1 1 plants ranged between 26 and 79% (molar-based) (29, Tables 2 and 4); the values in Table 13 of this paper indicate that between 30 and 75% elimination of total NPc during secondary treatment is achieved, or between 15 and 50% is discharged to the environment in secondary effluents. Differences in the relative importance of biodegradaton and sorption at each stage of sewage treatment for the four major nonylphenolic groups are depicted in Figure 15. Table 13. 96 Distribution of Nonylphenolic Compounds (Molar-Based)1 Raw infl 1 ° 2 ° Dig'd sldg Category effl effl 2 NPE3-20 57-63 55-62 3-104 "-5 NPELZ 28-30 15-18 2-12 1-2 NP 3-10 5-6 1-5 15-24 NPELZC <1-43 1-53 8-25 tr5 Tot'l NPc 5100 80-90 15-506 1 5-257 1 Unless stated otherwise, values estimated from 11 plant study; Ref. (29): text, Figs. 3 81 8, Tables 2, 4 & 8. 7- Ref. (29, Table 8 and Figure 8): Values estimated. 3 Ref. (15, Figure 12); slightly higher levels-of NPELZC measured in Uster 1° effluent than raw influent. 4 NPE9.20 not present; NPE3-g present in trace amounts. 5 Only trace levels present. 5 Discharged to surface waters. 7 Usually land applied, or land filled. 67 SEWER RT - mins. to hrs. "153-20: ll 8 WE 1,2: 118; is M’ : 118; 15 LP“ 12C: 13 PRIMARY SETTLING BIOLOGICAL TREATMENT RT - 10 days "’5 3-20 I 1111 3 M13 2 H S NP: us "£1.20 1111 B \/ Secondary . effluent - SLUDGE actsnon 5 9551339“ NAEROBIC/AEROBIC ; 1° 5“ e T - 30 days 1 waters ---------------------------------- Thickener f“ “1.221113/13 i NP: 1118/18 Wastewater ——> Sludge/solids --- up De' t ed RT - Residence time "a e" B - Biodegradation sludge — S - Adsorption Land applied Nurrber of arrows designates relative in‘portance of mechanism during total process, from minimal(one arrow) to major (four arrows). Figure 15. Relative importance of biodegradation and sorption as removal (I) and accumulation (f) mechanisms during water treatment - by major category. Primary treatment With a 2 hour retention time (Table 3) and high percent suspended solids in the primary tank, sorption is the major removal process, as depicted by the distribution curves for the Uster plant (29), Figure 13. The lowest and most lipophilic oligomers (NPE1&2) and NP, are considerably lower in the primary effluent than in the raw influent, down to approximately 100 umol/ m3 from 180 - 530 umol/m3. NPE3 is slightly lower, at 40 umol/ m3. The NPE4-13 oligomers change very little, demonstrating that biodegradation is not 68 a major removal mechanism in the primary treatment stage. This is illustrated in Figures 13 and 15. Secondary treatment The hydraulic retention time in the Ziirich-Glatt STP is about 6 hours (133); the mean cell residence time is not provided for any of the 11 plants studied, but the times provided in Table 3 are probably a close approximation (10 days). This allows ample time for biodegradation, the major removal process, to occur. The major findings are summarized below. 0 Nonylphenolic compounds are not the most biodegradable part of DOC; they are eliminated to a lesser degree than is overall BOD (70 a 1596 vs. 86 a: 996 (weight-based». Overall elimination efficiency of the NPEs and metabolic products during secondary treatment in the 11 plants studied by Ahel et a1. varied from 43 to 8796 (weight-based), and from 26 to 79% (molar-based) (29, Table 2). NPE3-20 is reduced substantially in the aeration tank, leaving only trace amounts of the NPE3-3 oligomers in the secondary effluent (Tables 12 and 13, Figure 14). NPE1&2C is the predominant product in secondary effluents (29, 1 16, 190). Formation of NPB1&2C during aerobic biological treatment occurs in all plants so that levels are 2.1 - 7.6 times higher in secondary than primary effluent (Tables 12 and 13, Figure 14). The apparent need for anaerobic conditions for transformation of NPELzC (13, 15, 29, 55) and a relatively low distribution coefficient (SOO L/kg for NPELZC vs. 10,500 L/kg for NP (Table 6)) contribute to its persistence and resulting high levels in secondary effluents. 69 o Persistence of NPE metabolites increases with increasing lipophilicity, or shortening of the E0 chain. The fate of the lipophilic products - NP, NPE1 and NPEz - is influenced by solid/ aqueous phase partitioning. The effect of formation of NP and N PE1&2 in the aqueous phase of secondary treatment is diminished by sorption due to much higher partitioning coefficients -- 10,500, 1800, and 900 L/kg measured in the Uster plant for NP, NPE1 and NPEz, respectively. Maintenance of a high suspended solids concentration (mixed liquor volatile suspended solids), typical in the activated sludge process (Table 4), facilitates the sorption of these compounds, thus removing them from the aqueous phase (Tables 12 and 13, Figures 14 and 15). 80er to sludge, they are further biodegraded during sludge digestion. If the sludge is later applied to agricultural land or placed in a landfill, any metabolites remaining may leach into groundwater, wash into surface waters as runoff, or be mobilized with colloidal material, perhaps increasing risk in additional environmental compartments. 0 Related to the above, the order of resistance to biodegradation is NP, NPEl, NPEz (most to least) (12, 126). Sorption, toxicity, and bioconcentration increase (NPEz < NPE1 < NP) as rates of biodegradation decrease (24-2 7, 5 4, 142). 70 o The elimination rates of products of NPE3-20 vary widely (Table 14). The highest elimination rates are seen in low-load plants. NPEl, 2 is formed more quickly than eliminated (-1996) in one high load plant. NPELZC is formed more quickly than eliminated (~660 to -110%) in all treatment plants. Elimination of NP varies from 9 to 9496. Table 14. Range of Product Elimination Efficiencies (Aqueous Phase) - Secondary Treatment‘ 96 Elimination - Range Product (Lowest to Highest) NPELZ ‘19 to 80 NPELZC -660 to -110 NP 9 to 94 1 Ref. (29); from study of 11 plants. 0 Lower temperatures result in lower overall elimination rates for NPEs in both lab and field experiments (Table 15). Studies that determine the effect of changing temperatures on biodegradability have been conducted in bench-scale systems simulating activated sludge treatment (1 7, 41), in a porous pot apparatus (36), and in the field in both biological filtration (36) and trickling filter systems (124). Naylor found only a 2-5% reduction in removal efficiencies due to seasonal effects, but specific temperatures are not provided (54). Biodegradability in natural environments cannot be predicted based on fluctuations in temperature alone. It is possible that organisms more tolerant of lower temperatures exist in some environmental compartments into which surfactants are discharged. In the event that acclimation has occurred, an increased population count or increased enzymatic activity of the degrading organism may result, and the biodegradation rate may actually increase when lower temperatures are introduced in a compartment ( 75) . 71 Table 15. Temperature Effects on the Extent of APE Removal Temperature Compound] System/ (° C) or Extent of lnfluent Conc. Test Method period Removal Ref. NPE9/ Bench-scale activated 15-25 a 92 (17) 10 mg/ L sludge/CTAS‘ NPE9/ Bench-scale activated 8 28 - 83 10 mg/L sludge/CTAS1 NPEg/ Bench-scale activated 25 > 93 (41) 50 mg/L sludge/ HPLC-FD2 NPE9/ Bench-scale activated 8 80 10 mg/L sludge/HPLC-FD2 OPE3-9/ Pilot-scale activated 10.5-14.0 79-88 (36) 7 mg/L sludge/Wickbold OPE3.9/ Porous pot activated 15 95 (36) 20 mg/L sludge/TLC3 OPE3-9/ Porous pot activated 1 ‘I 67-70 20 mg/ L sludge/TLC3 OPE8-9/ Porous pot activated 8 48-53 20 mg/L sludge/TLC3 OPE3-9/ Biological filtration/ March-May 20 (36) 10 mg/L Preston/TLC3 OPE3-9/ Biological filtration/ May-September up to 80 10 mg/L Preston/TLC3 OPE3-9/ Biological filtration/ September- down to 20 10 mg/L Preston/TLC3 December OPE3.9/ Biological filtration/ Nov '73 - Mar '74; 20-40 (36) 10 mg/L Langley/TLC3 7.6-10.0 OPE3-9/ Biological filtration/ Summer '74 80-90 10 mg/L Langley/TLC3 OPE3-9/ Biological filtration/ Nov '73 - Mar '74; ~ 80-90 10 mg/L Langley/TLC3 9.2-1 1.2 APE7,3 Trickling filter/ March '85 70 i 3 (124) 07710.06 mg/L HPLC—UV or FDZ APE7,5 Trickling filter/ Sept. '85 75 :i: 3 0.731006 m/L HPLC-UV or FDZ ‘ CT AS - Cobalt thiocyanate active substances. 2 HPLC~UV or FD - High pressure liquid chromatography-ultraviolet or fluorescence detection. 3 TLC - Thin layer chromatography. 72 Different conclusions are presented by Kubeck and Naylor (54, 131), who insist that: 1) the rate of biodegradation is not affected by the length of the ethoxylate chain (131); 2) all NPE1-13 oligomers are removed satisfactorily during sewage treatment (92.5 - 99.8%), even under high sludge loading conditions (54); 3) NP is not sorbed preferentially to sludge during treatment (54, 131). Research results that support the conclusions of Kubeck and Naylor for environmental samples could not be found. However, researchers working in laboratories around the world have published results and conclusions similar to those of Ahel, Giger, and co-workers: 1) Lower oligomers (especially NPEz and NPEZC) predominate after treatment of APEs and carboxylated APEs (9, 18, 21, 23, 41, 116); additionally, Swisher (6, pp. 721-724) cites the work of several researchers, in addition to Giger, who identified NPEz as an intermediate that accumulates when the higher Oligomer NPEs biodegrade ( 1 92-1 98); 2) NP and the lower oligomers are preferentially sorbed during primary and biological treatment (39, 128, 142, 194, 199). Closer examination of the results and conclusions published out of both laboratories follows. 0 In the published results of their 1990 study, Kubeck and Naylor conclude that NP and NPE1&2 do not accumulate during sewage treatment. They determine that improper work-up procedures in the research completed by W. Giger and co—workers cause the distribution of N PE to be skewed toward the lower oligomers. They identify the use of polyethylene storage 73 bottles, nitrogen or solvents contaminated with oxygen, interfering substances in the extract, or insufficient nitrogen streaming during blow- down as possible causes. They insist that NP and the lower oligomers (NPEl&2) are eliminated as efficiently as the higher oligomers. The data of overall elimination efficiencies, collected in May, 1988, from two plants (East and West) in High Point, North Carolina, and reported in their 1990 article, appear to support their conclusions (Table 16). However, their use of weight-based rather than molar-based measurements distorts their comparisons between higher and lower oligomers in terms of their reactivity (biological and chemical) (29). That is, distribution of the lower oligomers will appear to be less, relative to the higher oligomers, on a weight basis than on a molar basis. In spite of this, NPE Oligomer concentrations in the West High Point plant depict a distribution that is skewed toward the lower oligomers - NPB1, NPEz, and NPE3 (131, Figure 9). Table 16. Removal of NPE1-]8 in Two North Carolina STPs1 (pg/L) Date Raw infl Z' effl . . . Category (mo/yr) low/hiLh low/high 96ellmlnation NPEHg 5/88 16002/25203 513/1042 932/983 NPE1 5/88 363/692 2.63/34Z 933/95Z NP 5/88 nm4 0.83/25z - - ‘ Ref. (131); statistics not provided. 2 Eastside plant - industrial: textile processing and furniture manufacturing. 3 Westside plant — domestic. ‘1 Not measured. 0 Concentrations of individual oligomers measured at and near the Govalle, Texas plant, and published in the same article, also do not support their conclusion that the NPE Oligomer distribution is not skewed to the lower oligomers. On the contrary, graphs of effluent and river water data all depict elevated quantities of NPE1, NPEz, and NPE3 relative to the higher oligomers. The authors attribute the skewness to " the possibility of interferences", but no further explanation is provided. 74 0 Additional data, collected as part of a later study from High Point and six other plants, depict similarly high removals overall ( 5 4). Nine of the 14 results published by N aylor are displayed in Table 17 for illustrative purposes. Two mid-range results from each of the East and West High Point, NC plants, and an earlier sampling (8/ 90) from the Small City, Midwest USA plant are excluded. Naylor provides graphs of NPE Oligomer distributions for three of the fourteen treatment plant results, designated with '*' in Table 17. All three distributions reflect generally higher concentrations for NP and the lower oligomers - NPEl, NPEz, and NPE3 - to varying degrees (54, Figures 5-7). Table 17. Removal of NPE1-18 in Seven U. S. STPs1 019/ L) 96 in Plant (mga/ter) NI’in'1111'1)5 fl211.11..“ eliz'n dewat'd y sludge Burlington, NC (East) 5/92 384 + 903 14.4 + 54 94.6 Burlington, NC (South) * 5/92 359 + 2619 15.3 + 39 98.2 High Point, NC (East-high) 5/88 - + 1780 1.3 + 79 95.5 High Point, NC (East-low) 4/ 93 14 + 274 elim'n). . NP: 3796 elimination. 1 Ref. (29); comparison of two plants. Sludge digestion As discussed earlier, measurements of samples from digested sludge do not correlate with measurements of samples drawn from primary and secondary effluents. Measurements of NPE metabolite concentrations in digested sludge are not synchronized with the incoming load of nonylphenolic compounds. Average residence times can vary from two hours to ten days to 30 days for primary treatment, secondary treatment, and sludge digestion, respectively. Additionally, grab samples rather than composite samples are collected for analysis of digested sludge. Under these conditions, it is not possible to synchronize the withdrawal of samples from the various treatment stages. Though an attempt has been made in Table 13 to estimate the distribution of NPE metabolites in digested sludge relative to the primary and secondary treatment stages, it is important to keep this disparity in mind while studying the fate of NPE metabolites. According to the results of research completed by Giger, Ahel and co- workers, NP occurs in high concentrations in aerobically and anaerobically digested sludge of STPs that treat wastewater carrying nonylphenolic compounds. Its accumulation in digested sludge is a result of: 1) adsorption of NP and NPE1&2 to solids during primary and secondary treatment; 2) transformation of NPE1&2 to NP during sludge stabilization; 3) persistence of NP under anaerobic conditons. 78 Highlights of NPE research related to sludge digestion follow. 0 Partitioning of NP is controlled by the concentration of suspended solids -- the greater the concentration of suspended solids, the greater the partitioning (Table 19) (3 9, 128). The effect of the concentration of suspended solids in primary treatment is illustrated by the average percent removal range for total NPc that is estimated in Table 13 (10 - 2096). Higher removal, closer to 2096, will be realized when the suspended solids concentration during primary treatment is correspondingly high. Once the lipophilic metabolites -- NP, NPE1 and NPEz - are sorbed and settle during primary treatment, biodegradation ceases to be the major mechanism for their removal. They persist because with relatively high sorption coefficients, especially for NP, solid partitioning in the sludge will be favored over biodegradation. They may undergo further biodegradation during sludge digestion. Table 19. Relationship of Concentration of Suspended Solids to Fraction of NP Partitioning to Solids Concentration of Fraction of NP suspended solids 1 partitioning to solids 1'2 Medium (mg/L) (96) Digested sludge 50,000 ~ 100 Raw sludge 14,000 ~ 100 Activated sludge 1700 > 98 Raw sewage 87 ~ 10 Primary effluent 48 ~ 10 Secondary effluent 8 bd 3 1 Ref. (128). 2 Ref. (39). 3 Below detection. 79 0 Transformation of NPE1 ,2 to NP is favored under anaerobic conditions (1 3) . Brunner et al. confirm that more NP is present in sludges treated anaerobically than aerobically by measuring and comparing NP, NPE1 and NPEz concentrations in sludge samples from 29 SI‘Ps (128). Using 1986 data, the mean concentrations for NP were 1.27 and 0.30 g/kg dry matter in anaerobic (24 plants) and aerobic (5 plants) sludge samples, respectively. Furthermore, levels of NPE1 and NPEz were lower in plants that treat sludge anaerobically. NPE1 and NPEz mean concentrations in anaerobic samples were 0.19 x 0.12 and 0.06 a: 0.04 g/kg dry matter, respectively, vs. 0.36 a: 0.28 and 0.13 x 0.11 g/kg dry matter, respectively, in aerobic samples. 0 Once formed, NP persists, remaining sorbed unless exposed again to aerobic conditions. Data developed by Ahel et al. (29) show that more than 90% of the NP that is discharged to the environment is associated with digested sludge. From another perspective, the extrapolation of studies in the Uster and Ziirich-Glatt plants indicates that a relatively large fraction (15 - 25%) of the total nonylphenolic compounds that enter an STP leave the plant in the digested sludge, a large part of which is NP (Table 13). o Reintroduction of sludge to an aerobic environment may occur following stabilization, with sludge disposal. If the sludge is land applied or landfilled, NP may biodegrade if conditions are favorable. If NP is desorbed, it may infiltrate or leach into groundwater or enter surface waters as run—off (Figure 3). Alternatively, NP may remain sorbed to solids (colloids), and in this way become mobilized. Research on the effects of colloids is in its infancy and the body of evidence is relatively small, emphasizing the need for more studies that will ascertain the environmental fate and behavior of NP as a result of sludge disposal. H) Contrary to the conclusions of Ahel et al., Naylor (54) concludes that metabolites of NPE do not accumulate in sludge and are not an environmental risk. Comments about Naylor's research, as related to the sorption of NPE metabolites in sludge, follow: 0 In his 1995 study, Naylor provides the percent of NPE metabolites sorbed to sludge for only two plants (54). Results are reported in Table 17. More measurements are required to validate these readings. 0 Measurement of NPE metabolites in dewatered sludge yields very low values, 0.1% and 1.7% (predominantly NPEz) of total influent load, compared to 20 - 25% as determined in research published by Ahel et al. (29, Table 8). This implies much lower partitioning coefficients for NP and the lower oligomers than indicated in other research (39, 128, 142). A possible explanation for the low percent sorption values is experimental error. If NPE metabolites are not desorbed from sludge or glassware during extraction, then the amount of NPE partitioning to sludge will be underestimated. W Table 20 provides results of several research studies that have been completed worldwide to determine levels of NPE and metabolites in raw, mechanically treated, and biologically treated effluents. Figure 16 is a bar chart of Table 20 values. Measurements from the Ahel et al. (29) study are included, with weight-based rather than molar-based concentration values. The data provides a comparison of measurements from varying parts of the world. Environmental levels cannot be directly compared because test conditions, sampling methodology, analyte measurements and analysis techniques are different (27). 81 Table 20. Concentration of NPE and Its Metabolites in STPs (ug/ L) Location 111153-20 NPEuz NP NPEnzC Analysis] (istd dev) (tstd dev) (:tstd dev) (istd dev) Ref. Switzerland (NE) HPLC‘1/ (29) 11 n . I tt R. 1° effluent 1181 173 43 43 2° effluent 112 88 16 184 96 elimination 90.5 49.1 62.8 -327.9 Switzerland (W) SlM-GC/ M55 Mm /(14) Vidy STP 2° effluent - - 387.8 35.8 - - Nyon STP 2° effluent - - 61.5 44.6 - - Prangins STP 2° effluent - - 106.1 69.3 - - It a ly LCs/(175) SEE: Raw influent 213.31 18.6 - - (11:10.8) (21:11.3) 2° effluent 10.31 1.8 39.6 2 (17.5) (i1.0) (21:35.2) 96 elimination 95.2 90.3 pp 3 U S A PB-LC/MS7/ New Jersey (184) Plant A - 2° effluent 408 - - - - - - Plant 8 - 2° effluent 139 - - - - - - Palo Alto TP OPE] ,2 OPEC3-4 OPEC“; (111813” 5 day average 4.5 11 8.0 11 56.1 (12.0) (12.8) (325.7) 1 Includes NPE3-20 7- NPE1&Z&3C; NPEmzC- 22 (i: 10) ug/L (quantitative); NPE,3C - 18 (:I: 28) [Lg/L (semi-quantitative). 3 Biodegradation product: formation > elimination. ‘High pressure liquid chromatography; 1 1 plant average of 24-hour composite samples (statistics not provided). 5 Selected ion monitoring gas chromatography/ mass spectrometry; mean values of two determinations (statistics not provided). 5 Liquid chromatography; annual average of 12 monthly 24-hour composite samples. 7 Particle beam-liquid chromatography/mass spectrometry (statistics not provided). 3 Semi-quantitative; NPE... 9 Semi-quantitative; NPE4 = 1 ug/L; NPE5 =- 12 ug/L. 1° Gas chromatography/ mass spectrometry; 5 day average of flow-weighted composites from three grab samples per day, @ 9am, 2 and 7 pm. 11 Value inflated because 0.5 pg/L used for three readings that were below detection level of 0.5 ug/L. [1]]NPE3.20 DNPE12 INP ENPELzC A 400 3“ u Ill NE Swtz- W Swtz- W Swtz- W Swtz- ltaly- Glatt Lk Geneva Lk Geneva Lk Geneva Ostia River 1 2 3 USA-New USA-New Jersey A Jersey B Figure 16. Concentrations of NPE and metabolites in secondary effluents. Values from Table 20. Wrens As is the case with sewage treatment plants, research studies address the behavior and fate of the APE surfactants in terms of the four categories defined previously - APE3-20, APEl 5,2, AP, APE1 5,2 C. Again, biological degradation and solid/ aqueous phase partitioning are the major processes that determine the behavior and fate of nonylphenolic compounds in river systems (water column and sediment). Hydrodynamic conditions (longitudinal and transverse dispersion) also must be considered (133). Volatilization (1 99, 202) and photochemical reactions (133, 203) may play lesser roles in reducing environmental levels of surfactant metabolites. A review of the major findings about NPE behavior and fate in surface waters, with emphasis on the Ahel et al. (1 33) river study, is provided below. For the second part of their study, Ahel et a1. drew secondary effluent samples and river water samples (one day, 12 successive Z-hr. composite samples) at several stations along the Glatt River during February (winter) and August (summer), 1984. Sampling times were set so as to obtain the same water package; STP and river sampling were synchronized. During other seasons grab samples of river water and sediment were taken from different locations along the river (133). Water column The elimination process that begins in the STP continues in the river. Less than 596 of total NPc enters the river at its inlet, where Lake Greifensee flows into the Glatt River. The remaining load is introduced in the secondary effluent that is discharged to the river from STPs. Almost 88% of total NPc load is released to the Glatt River within the first 15 km of its 35 km length, and most of this is released between 12 and 14 km from the inlet (101). Measure- ments of NPE and metabolites in river waters are provided in Tables 2 1 84 (mol/day) and 22 (ug/ L). The average hydraulic retention time of the majority (75%) of total NPc input to the Glatt River is estimated to be between 10 and 15 hours, long enough to effect biological transformation (affected by light intensity and duration, amount of rainfall, and temperature) and physical- chemical processes. An estimated 15 - 20% of the water at the outlet of the river is treated wastewater (133 ). o Ahel et al. found that 24% overall elimination (molar flow) of N PE and metabolites occurs in the river along the 35 km stretch studied (Table 21). This falls below the range of elimination efficiencies that were measured in the Glatt River STPs (26 - 7996). 0 Similarly to secondary effluents, the presence of NPE3-20 oligomers is minimal. Biodegradation is the major process by which the parent surfactant is removed from the water column. Bio-transformation of the parent surfactant, NPE3-20, and its products, NPE1&2, to the carboxylated form occurs rapidly in the aerobic environment of river water during transport, according to the pathway: NPE3-20 '*’ NPELZ —’ NPE1,2C. The distribution of total NPc introduced in the secondary effluent changes in favor of the carboxylated products -- NPE1,2C - up from 51 to 86%, while the parent compound, NPE3-20, and the noncarboxylated metabolites, NPE1&2 and NP, decrease to 3, 9, and 296, respectively (Table 21, Figure 17). NPE1&2C persist. 85 Table 21. Inflow and Outflow of NPEs in the Glatt River Inflow Outflow Elimination Inflow Outflow Component (2' effl) Distribution Distribution (moI/d) (mol/d) (96) (96) (96) NPE3-20 23.4 2.8 88 22 3 NPELZ 24.3 7.2 70 22 9 NP 5.3 2.0 62 S 2 NPELZC 55.2 70.1 -27 S1 86 NPc 108.2 82 24 100 100 Ref. (133). Secondary Effluents m Outflow Input - 108.2 mol/day Output - 82 mol/day 22% DNPE3.20 .NPE 51% 1'2 .NP .NPELZC Figure 17. Distribution of NPE compounds in secondary effluents and Glatt River water. Adapted from (133). 86 0 large fluctuations in concentrations of NPE and metabolites occur along the river (Table 22), varying with location, season, and time of day. The high measurements for NPE1&2 (69 rig/L) and NP (45 [lg/L) are outliers; of the samples measured, 219 and 109 measurements fall in the range 0 - 30 [Ag/L and O - 8 [Lg/L, respectively. Table 22. Environmental Analyses of NPEs in River Water and Sediment -— by Major Category 1 River water River sediment Component Number of low/high low/high samples (jig/L - ppb) (uflg - ppm) 2 NPE3-20 «3 <1 ./ 7.1 --‘1 NPELZ 220 <0.3/695 . 0.29/11.25 NP 110 <0.3/4515 0.51/13.1 NPE1.2C 96 <1/71 "4 NPc NA 1.27/24.35 1 Ref. (133). 2 Based on dry weight of sample. 3 Not provided. ‘1 Not measured. 5 One NPE, sample was 69 rig/L; 219 samples were < 30 [Lg/L. 5 One NP sample was 45 rig/L; 109 samples were s 8119/ L. 0 Longitudinal profiles based on product concentrations vs. mass flows (both winter and summer data) depict the role that dilution plays, together with biodegradation and physical-chemical processes, in decreasing aquatic concentration levels. - Concentrations of all NPE products increase with the input of secondary effluents up to the test station located 15 km downstream. Between the test stations that are 15 and 35 km downstream, the concentrations of all products decrease, Figure 18. -- Mass flows also increase up to the test station that is 15 km downstream, but they are more likely to remain relatively flat or increase between the 15 and 35 km test stations, Figure 18. 0 The effect of improved efficiencies of elimination that occur in the STPs during warmer weather are reflected in the seasonal longitudinal profiles 87 measured from the river water, Figure 18. River water temperatures during winter and summer samplings are 4.8 - 5.6 and 18 - 20°C, respectively. Summer readings of the NPE product concentration levels and mass flows are generally lower at all downstream test stations. Improved elimination efficiencies in the river water due to warmer temperatures are also evident. Dilution attributed to seasonal variations in river flow affects concentration levels minimally, if at all. - NP: Data measurements are slightly higher during the cold season, Figure 18. NP remains relatively flat during both seasons, probably because sorption, a process that is much less sensitive to temperature, is the mechanism for its removal rather than biodegradation. -- NPE1,2: Winter measurements are 2 - 7 times higher than summer measurements. The winter mass flow increases about 1596 in the last 20 km stretch of the river, while the summer mass flow decreases 30%, Figure 18. -— NPELZC: A longitudinal profile for NPE1,2C is not provided for the winter months. Summer longitudinal profiles portray a 1596 increase in mass flows beyond the 15 km point, probably due to higher biodegradation rates associated with warmer temperatures, Figure 18. o Nonylphenol ethoxylates and their derivatives are the predominant organic compounds in the Glatt River system, exceeding other hydrocarbons (HCs), chlorinated HC5 and chlorinated phenols (12). Concentrations of the individual oligomers of nonylphenolic compounds are one order of magnitude higher than other compounds, such as tetra- chloroethylene, 1,4-dichloro-benzene, 1,3-dimethylbenzene and pentachlorophenol, measured in another study ( 65). .Amm: :3. so: 6883‘ .82 35.69: .28 ecu mozugtou $2 60530:. we Amie: mmmE ecu 8:828:08 $.an 3:635:04 .2 950.1.— E85238 Ex eucaama Enobmssou Ex 3:320 az ~+Ez o 8132 w M £2 .23 o. W, m M m I II 8 Om. 0 w 0m Ewwhumcgov Sufism—030$ Ex 8:220 Ev. 8:390 ~+Ez 8232 m. m. 22 .38 w w m .m w w .353 Water column and sediment Partitioning to organic material, suspended solids, or colloids, and subsequent settling remove NP and NPELz from the water column. Photochemical degradation (203) and volatilization (1 99, 202) may be removal mechanisms in surface waters that have very little organic material, or very low concentrations of suspended solids or colloidal material. However, the high concentration of colloids, suspended solids, and organic material, typical of many surfactant relevant environmental compartments (Table 4) favors sorption over photolysis and volatilization. Although the fraction of organic material, and concentrations of colloidal material and suspended solids along the Glatt River are not reported, Ahel et al. report the following: 0 Higher concentrations of NP, NPE1, and NPEz are measured in mud than in sand collected at the same location (133). Partitioning is greater in the mud because of its higher organic content. The concentration levels in Table 22 provide evidence that the concentration of NP and the lower oligomers (NPE1, 2) is about three orders of magnitude greater in Glatt River sediment than in the river water. 0 Algae growth is so thick in parts of the Glatt River that it is removed mechanically at the end of every summer (26). Ahel et al. have measured concentrations of NP, NPE] and NPEz in algae as high as 25, 80, and 29 ppm (mg/kg), respectively. 0 Not only does the algae offer an environment onto which NP and the lower NPE oligomers can sorb, but its accumulation leads further to a disposal problem when it is removed annually from the river (26). 90 0 Longitudinal profiles of NP, NPE1 and NPEz in river water and sediment are similar (Figures 18 and 19). The sediment concentration levels of all three metabolites decrease between the 15 and the 35 km sediment test stations. NPEl, 2C values are not provided for sediment. o NP persists under anaerobic conditions prevalent in sediments. It remains sorbed to organic material in the sediment unless turbulence (fast water flow or bioturbation) effects resuspension or aerobic conditions, in which case it may biodegrade. The extent to which NP accumulates needs to be _ determined. Ahel et al. measured the ratio of NP in sediment to NP in water to be between 364 and 5100 (133). —D—NP Concentration (ug Ig-ppm) 0T . i a 1km 9km 15km 35km Distance downstream Figure 19. Longitudinal concentration profiles of NP, NPE], and NPEZ in sediment. Adapted from Ref. ( 133). 91 Wr -Srf ter Tables 23 and 24 and Figures 20 and 2 1 provide information about NPc concentrations that have been measured for surface water samples collected worldwide. The data cannot be directly compared because test conditions, sampling methodology, analyte measurements and analysis techniques are different (2 7) . The studies reported in Table 23 were completed in Switzerland and the United States, in fresh water environments. A wide range of aqueous concentrations is evident for NP (Table 23), from as high as 3000 ppb in a small creek that receives the overflow from an anaerobic treatment pond of a carpet yarn mill (136) to as low as below detection levels (0.11 ppb) (54). In the former study, NP concentrations in samples drawn from several downstream locations decrease rapidly as distance from the treatment pond increases, so that only 2 ppb NP is measured in a Savannah River sample taken 9.6 km from point of discharge. As discussed above, this is due primarily to hydrodynamic conditions. dilution, and sorption. Concentrations in sediment samples also vary (Tables 22 and 23), ranging from 13,100 ppb (133) to below detection levels (2.99 ppb) (54) in samples measured in Switzerland and the United States, respectively. 92 Table 23. Environmental Analyses of NPcs in Fresh Water and Sediment 1-2 C Analysis/ Location NPE3-20 NPEjaz NP NPEtaz Ref Switzerland filatt Rivgr HPLC/( 133) Low/high <1. /7.1 <0.3/69 <0.3/4S <1./71 Switzerland Distance/ Lk Geneva depth SIM-GC/ Average (n - 2) (m) MS/(14) e Vidy 200/ 6 --3 9.7 2.9 --3 @ Nyon 5-10/4-5 --3 2.4 1.2 --3 @ Prangins 5-10/4-5 --3 8.3 3.4 --3 USA GC/MS/FID/ fieorgia ‘1 Distance 5 (km) (136)‘1 Discharge 5 --3 3 4000 --3 Small creek 1.6 --3 3 3000 --3 Small river 2.4 «3 --3 200 --3 3 3 3 3 3 3 Small river 7.2 -- Savannah River 9.6 -- USA lhirty Livers study-water HPLC/(54) Highest 14.9 1.8 0.64 --3 Average (n - 100) 2.0 0.2 0.12 7 --3 Thirty rivers study-sediment (ml kg) (119] kg) HPLC/ (54) Highest --3 1758 2960 --3 Average (n - 80) --3 188 1629 --3 1 Data reported in [lg/L unless designated otherwise. 2 Statistics not provided. 3 Not measured. 4 NP concentrations in textile waste and receiving waters. 5 Downstream from textile mill's anaerobic treatment pond discharge. 5 From textile mill's anaerobic treatment pond. 7 NP was below detection (0.11 ppb) in more than 7096 of samples. 3 NPE] only measured. 9 NP was below detection (2.99 ppb) in 28% of samples. Table 24 provides data from research that examines NPc behavior and fate in marine environments. The data is collected from two locations that border the Adriatic Sea, in Croatia (137) and Italy (138). Again the range of concentrations that are measured in the two studies span up to three orders of magnitude. The Venice lagoon receives treated wastewaters from domestic and industrial sources. This is evident in the lower concentrations for NPE1-13 and 93 its derivatives - NP, NPE1&2. Concentrations in water, associated with RSM (resuspended solid material), and in sediment cores are reported in Table 24. Croatia discharges untreated municipal wastewater into Sibenik Harbor, as reflected by the very high average concentrations of NPEp (p=0-20) released from four sewer line outlets, Table 24. Analysis of wastewater samples yields NPE oligomeric distributions that are characterisic of NPE found in heavy duty detergents; the maximum of the oligomeric distribution is at 10 E0 units/NPE molecule. Partitioning followed by sedimentation and dilution are the major processes by which NPEs are removed from the water column closest to the sewer outlets. The effect of partitioning is examined by measuring the NPE concentrations of dissolved vs. particulate fractions in the sewer lines and at the sewer outlets. A significant amount, up to 6096 in some samples, is removed from the water column when it rapidly sorbs to particulate matter in the sewer line. Figure 20a is a pictorial representation of the change in total concentration and the dissolved and particulate fractions as the wastewater moves through the sewer line to its point of discharge. The sewer line and outlet concentrations of NPE associated with suspended solids based on the dry weight of suspended solids are also provided in Table 24, and are illustrated in Figure 20b. The sewer line and outlet concentrations of NPE based on the dry weight of suspended solids are very similar, indicating that desorption is either not occurring or is negligible. Additionally, oligomeric distributions of the dissolved and particulate fractions demonstrate the importance of lipophilic partitioning in determining behavior and fate of NPEs in wastewaters. The oligomeric distribution of the dissolved fraction is similar to the total sample, but the central and maximum point is shifted to the higher oligomers; oligomers with between six and eight EO units/NPE molecule are predominant in the particulate fraction. 94 The decrease observed in total concentrations at the sewer outlet is the result of dilution upon mixing with marine water in the harbor. The effects of further dilution are evident in measurements of estuarial NPE concentrations, which decrease exponentially as distance from the outlet increases from 1 to 200 m, Table 24. Concentrations of suspended solids drop from 122 - 884 mg/ L in the wastewater to between 1 and 5 mg/L in the estuary, ruling out sorption as a major process in removing additional NPE from the estuarial water column. 95 Table 24. Analyses of NPcs in Marine Environments - Aqueous/Solid Partitioning Dissolved Particulate Total Dry weight fraction fraction conc'n conc'n Distance (pg/Lied) (pg/Ltsd) (pg/Lisd) (mg/kg) I depth low/high low/high low/high low/high (m/m) Croatia1 Sibenik Harbor, Kgg R estuag 7 municipal W "pep 3711231 20911 15 5801290 8711468 127/783 50/406 177/1189 300/1500 4 gagglets2 NPEp 2521146 50:31 3021177 8501574 124/408 16/83 140/491 500/1700 In estan W593 --10 --10 126192 «10 1/0.5 43/257 NPE”4 --1 0 --10 0.4102 --1 0 25-200/ <0.1/0.7 0.5 NPED“ --1 0 --1 0 0210.1 --10 25-200/ <0.1/0.4 6.0 Conc'n in Conc'n on Conc'n in water-ppb RSMG-ppm sediment-ppb (pg/Lisd) (pg/gisd) (ng/gtsd) low/high low/high low/high Italy 5 Venice lagmn NPE]-]37 1.83:1 "10 "1° “10 0.5/4.5 NPE1&2 "1 0 1.4128 "1 0 351249 0.2/8.1 13/91 NP --1 ‘1 0.89:13 --1 0 14193 0.1/5.6 5/42 1 Ref. (13 7); analysis - HPLC (normal- and reversed-phase). 2 March 1991 samples; untreated water. 3 July and August 1990 samples; measurements from two outlets. 4 July and August 1990 samples. 5 Ref. (138); analysis - HPLC; data collected during Febniary, April, July, and October 1987. 3 RSM: resuspended solid material. 7 5 stations measured during 3 sampling sessions; April, July, October 1987. 3 Based on dry weight of solid material; 5 stations measured during 8 sampling sessions; February, April, July, October 1987, each at low and high shear stress. 9 Based on dry weight of solid material; 5 stations measured during 4 sampling sessions; sediment core samples - 5 cm. 1° Not measured. (a) (92 A 6w 3 10m «:1 500 g 800 g1 400 E g c 600 3 300 --‘-’ 40 mm 4,; 400 *2 200 .2 0 G g 100 g 200 0 o - . ° 0 0 III In D "' 8| '0' 5 g E a e 5 ti 2' 5 o S 3 o .1 8 to 0 0‘) D dissolved particulate dry weight Figure 20. NPE1-20 marine environment concentrations, from untreated sewer waters discharged to Sibenik Harbor, Croatia: (a) in dissolved and particulate fractions based on water sample volume, and (b) based on dry weight of suspended solids. Adapted from (137). Figure 21 depicts the effect that stratification in the water column has on NPEp concentrations in Sibenik Harbor, at some distance 25m or more from the sewer outlets. The highest concentrations are observed at the air-water surface microlayer and the brackish water-sea water interface. The interface between the brackish and saline layers serves as a barrier to vertical transport; higher NPEp concentrations within its boundaries are due partly to its high organic content. Wind currents explain the range of depths possible for the interface, and the overlap in the depth variations of the top and bottom saline layers. The top layer (0 - 6 m) is brackish, and has higher NPEp concentration values (1.1 - 6 ug/ L). The NPEp concentration in the bottom, saline layer (0.2 - 40 m) is much lower (<0.1 - 0.7 ug/ L). Two determinations provide evidence that biodegradation occurs in the estuary waters, although at lower rates than in fresh water. Warmer temperatures favor biodegradation. Winter and summer samplings indicate that the most abundant oligomers in 97 estuary water are NPE10.11 and NPE-1-9, respectively. Additionally, the proportion of NPE1&2 relative to total NPEp concentration is higher in estuary water than in wastewater samples. The authors conclude that biodegradation, rather than sorption, is the more likely removal process in the estuary because at less than 5 mg/ L the concentration of particles does not favor partitioning. Volatilization and photolysis may be minor removal processes that the authors did not address. (33 ' 17 Fig/L) Surface microlayer 0.2 - 0.4 urn Brackish layer Salinity-8-3096 0-6m (1.1 - 6 rig/L) Interface depth V Salinity - 8-32% {718 - 9 ug/L) 0'2 ' 6 m Saline layer Salinity - 35 - 40% 1‘0" ’ 0'7 "9’” 0.2 - 40 m @ Sediment boundary (<0.1 - 2.6 jig/L) 40 m Figure 21. Concentrations of NPE9 in the vertical profile of Sibenik Harbor, Croatia. Wis; Ahel et al. chose two sites for the third phase of their study, involving the Glatt and Sitter Rivers in northeastern Switzerland. Their study addresses the fate and behavior of NPE metabolites in aquifers with relatively short residence periods (62). The aquifers receive water that infiltrates through a river bed. Relatively high velocities - Glatt: 0.1 - 0.3 m/hr (65); Sitter: 4 m/hr during pumping (62) -- account for the short residence times. The NP and NPE1&2 data were collected over a period of 14 months (July, 1984 to August, 1985) at approximately monthly intervals; up to seventeen samples were collected. Only two samples of NPE1&2C were collected, on May 14 and 20, 1986. Samples of river water and groundwater were drawn from observation wells placed at 2.5, S, 7, 13, and 130 m intervals along the groundwater flow line (65). The plunger pump used to draw the samples was placed approximately 0.5 m below the water table to ensure that infiltrated Glatt River water was analyzed (65 ). Figure 22 provides a 0 - 13 m longitudinal concentration profile of average concentrations of the samples gathered from the Glatt River site for each of the five metabolites. Table 25 is a tabular version of the data plotted in Figure 22. The values in Table 25 are in terms of percent elimination; approximate residence times that correspond to the first four observation wells are also provided for minimum and maximum velocities measured in the aquifer. Samples were drawn from the Sitter River site to determine whether variations observed in pollutant concentrations are similar in river water and groundwater. Grab samples were collected in September, 1985, at two hour intervals for 24 hours. NPE1&2C are not measured. Analysis of sample measurements from the Sitter River site confirm the findings from average 99 measurements in the Glatt River study. The findings are summarized as follows: . o NP, NPE1 and NPEz are eliminated efficiently from the aqueous phase, on the average. Most of the elimination occurs within the first 2.5 m of the river bed (Figure 22 and Table 25 ). High organic matter and probably a higher population of microorganisms both contribute to the removal efficiency. - NPE1C and NPEZC are eliminated more slowly and to a lesser extent -- both within the first 2.5 m from the river bed and over the total distance (Figure 22). NPE1C and NPEZC continue to be the predominant metabolites of NPE in groundwater. They appear to be more mobile than the other NPE products. -- NPE1C and NPEZC are more soluble yet more resistant to biodegradation than the noncarboxylated oligomers. -- The difference is even more remarkable if one considers that the averages for mac and NPEZC are based on two samples collected during May, 1986, a warmer season. If cold weather results had been included in the calculations the inputs (measurements in river water at O m) would have been higher. More studies are necessary to determine how colder temperatures affect the rate and extent of elimination efficiencies, though it is likely that slower rates and lower elimination efficiencies would prevail. 0 Dilution also plays a role in eliminating NPE metabolites from this aquifer, owing to its relatively high velocities. 0 Order of elimination efficiency from highest to lowest is: NPEz z NPE1> NP> NPEzC > NPE1C 25 C o n 20-- c e nulSfl t9 r/ aL t 10$ 1 o n s 0 0 5 1(1) 10 Distance from river bed (m) —I— NPEZC —-0—NP —-1t— NPEZ —-o— NPE, Figure 22. Longitudinal profile of average concentrations measured in Glatt River and underlying aquifer, at 0.5 m below the water table. Adapted from (62) Table 25. NPc Elimination in Groundwater Underlying the Glatt River Distance from river bed (m) 2.5 5 7 13 Component 96 Elimination NPEZC 1 9.3 38.1 41.3 79.4 NPEIC 1 25.9 34.0 80.3 69.4 NP 2 64.4 85.2 83.7 92.6 NPE;z 96.1 98.1 99.3 >99.9 NPE} 2 88.3 97.7 98.8 99.5 Approximate residence time at above distance (daySl vel - 2.4m/day >1 > 2 < 3 5-6 vel - 7.2m/day > 0.3 < 0.7 < 1 < 2 1 Based on 2 samples. 2 Based on 16 samples. 101 0 Concentration measurements of NP, NPE1 and NPEz in winter samples of river water are 2 - 4 times higher than those taken during warmer seasons. Figure 23 provides NP, NPE1 and NPEz concentration levels that were determined in river and observation well samples during October, 1984, and January and June, 1985. Throughout the winter season, elimination efficiencies of NP and NPE1 are lower at the 2.5 m observation well. According to the January, 1985 data, the elimination efficiency of NPEz does not appear to be affected in winter at the 2.5 m observation point. Seasonal variations are mixed for the metabolites at the S m well. In general, the metabolites are eliminated more efficiently during the warmer months. This illustrates the importance of biotransformation as an elimination mechanism, and the role that temperature plays in affecting biodegradation rates. A range in temperatures (4 - 20°C) was only detected within the first seven meters of the aquifer. Seasonal variations in temperature did not affect elimination efficiencies for any of the metabolites at the 7, 13 and 130 m wells, where the temperature is fairly constant at 10 - 12°C. 102 NP 20 3 'g A 15 L A 2» 1° .. o .‘ '.'.'i 2 0 2.5 5 7 13 Distance from river bed (m) NPEI .5 i a C 0' 3 .3 5 O 0. 4.3:. 0 2.5 5 7 13 Distance from river bed (m) NPE2 : 20 O 2:: 15 2 Q E Jan-85 *5 u. 10 E 3 5 Oct-84 O 0 o I] Jun-85 0 2.5 5 7 13 Distance from river bed (m) Figure 23. Seasonal variations in NP, NPE1, and NPE; concentrations, measured in river water and groundwater, 0.5m below water table. 103 0 Wide fluctuations in concentration levels of NP and NPEz occur during the winter months. The data is provided in Table 26 and portrayed in Figure 24. February, 1985 measurements of NP are high —- 29 and 33 rig/L at the 2.5 and 13 m wells, respectively. A high NPEz measurement of 23 ug/L, from December, 1984, is depicted for the well that is S m downstream. More knowledge about biological transformations and transport in aquifers is necessary to ascertain the reasons for the wide variations. The authors provide two possible explanations: - A highly polluted "package" of water might move into the river bed such that the concentration of pollutant exceeds the ability of the surrounding aquifer to adsorb it and/ or biotransform it. Breakthrough of the pollutant would cause a huge increase in concentration. -- The likelihood that anaerobic conditions might establish within the first few meters of the aquifer could cause rapid biotransformation of NPE1&2 to NP. Table 26. Fluctuations in NPE Product Concentrations - High Readings from the Glatt River and Underlying Aquifer (pg/L) Product 0 m 2.5 m 5 m 7 m 13 m 130 m NP 26 29 I 4.4 3.4 33 1 7.2 NPE1 20.0 4.8 4.9 4.4 1.7 3.4 NPE; 21 1.6 23 1 0.7 0.1 1.7 1 Extraordinarily high readings. 104 Highest Winter Readings NW OM 0910 N U! oncentration ( J z ‘0 (Duo N O 0 S 10 15 Distance from river bed (m) Figure 24. NP and NPE; concentrations fluctuate in aquifer during winter months. 0 The extent to which sorption affects transport and ultimately elimination in groundwater can be estimated by considering the retardation factor: Rf = 1 + 3.2%M0c (Kow)0-72 p(1-e)/c It is calculated as a function of organic matter and the octanol/ water partition coefficient (62, 65, 100). To obtain low and high end retardation values within S m of the river bed, Ahel et al. assumed that the fraction of organic carbon in the aquifer (foc) is 0.001 - 0.01, and that the fraction of grain size < 125 mm (1‘) is 0.2 - 0.4. Density of the aquifer material, p, is 2.5 g/cm3 and total porosity, e, is 0.2. Using log Kow values equal to 4.48 and 4.12 for NP and NPE1&2, respectively, Ahel et a1. obtained retardation factors ranging from 11.7 to 216 (NP) and 6.9 to 119 (NPE1&2). If log Kow is equal to 4.2 (the approximate value in Table 6 for NPE1&2) (100), the estimated retardation factor range is between 7.8 and 136. This is still at least 63% below the range calculated for NP. Ahel et al. (62) suggest that biotransformation of NPE1&2 to NP explains the apparent greater transport capabilities of NP over NPE1&2 in spite of a retardation factor for NP that is higher by a factor of 1.5 to 1.8 (62). Transport of colloidal material associated with NP may also explain its apparent mobility. 105 0 High retardation factors extend the length of time that NP and NPE1&2 remain in the portion of the aquifer closest to the river bed. This enhances opportunities for biotransformation, while also keeping the pollutants close to the surface of the river bed. Whether biotransformed or not, they may reenter the river water by resuspension or desorb and be transported with groundwater flow (Figure 3). Concerns that nonylphenolic compounds may adversely affect groundwater quality are justified in light of data collected at pumping stations located long distances from NPc sources. Water samples from a pumping station located 130 m from the Glatt River provide evidence that NP and NPE1&2 have not been completely removed (Table 26) (62). NPE1&2C was not measured. A study, conducted at three field sites of an aquifer that receives infiltrated water from the Sava River near Zagreb, Croatia, also yields detectable quantities of NPE derivatives - NP, NPE1&2, and NPEC (Table 27) (68). The reader will note that the values are much smaller than those recorded in Table 26. They are included for the sake of comparison, and to emphasize that the carboxylated NPE derivatives, according to the data collected from Sites 1 and 2, persist, and therefore demand further study. Likewise, NPE1&2C was detected (not quantified) during analysis of sewage contaminated groundwater from a site on Cape Cod, Massachusetts (63). Table 27. NPc Concentrations in Groundwater in Zagreb, Croatia (pg/L)1 Site 1 Site 2 Site 3 Sava Waste- Product River 120 m Landfill 30 m water 15 m NP 0.70 <0.10 <0.10 --2 10.3 0.28 NPE1 0.40 <0.10 <0.10 "2 34.0 0.03 NPE; 0.20 <0.10 <0.10 --2 1.90 0.01 NPEC 5.00 0.20 0.05 0.05 0.29 0.004 ‘ Taken from (68, Table 4). 2 Not determined. 106 Summary. Primary biodegradation of alkylphenol polyethoxylates (APE 3- 2 0) generally proceeds rapidly in sewage treatment plants. Lower temperatures reduce elimination efficiencies during the winter months. Biodegradation proceeds by step-by-step elimination of the hydrophilic ethoxy units, down to two or fewer units (AP and APE1 5,2) . Under aerobic conditions, common in activated sludge processes, carboxylation occurs on the a-carbon of the terminal ethoxy unit such that APE1&2C are produced and accumulate. High quantities of carboxylated alkylphenol ethoxylates (8 - 2596 of total introduced APEp), and lesser quantities of APE3-3 and APELz (up to approximately 10 and 12% each, respectively, of total introduced APEp) are released to surface waters in secondary effluents. AP and APE1&2 that are not carboxylated or released in secondary effluents sorb to suspended solids and settle with the sludge in the primary or secondary clarifier (Figure 4). APE1&2 are biodegraded to AP during sludge digestion; biotransformation of APE1&2 to AP is favored under anaerobic conditions. Between 15 and 2596 of alkylphenolic compounds introduced to an STP are released to the environment in digested sludge. Alkylphenolic compounds remaining in the digested sludge are AP (approximately 95% or 15 - 24% of the total APE introduced) and APE1&2 (5% or 1 - 2% of the total APE introduced). Land application of the digested sludge may initiate further breakdown of AP under aerobic conditons, or lead to mobilization of AP that is associated with colloidal material. The elimination of APE and its metabolites from surface waters must be studied in terms of their behavior in association with particles and sediment, as well as in the water column. Biotransformation, sorption and dilution are the primary processes at work in the water column. Volatilization and photolysis are less important. Higher biodegradation rates that occur with 107 higher temperatures result in lower concentrations overall during summer months; lower concentrations are introduced to surface waters and observed in surface waters downstream of discharge points due to greater elimination efficiencies. The biotransformation of APE3-20 to APE1 3,2 to APB1&2C occurs under aerobic conditions in the water column. Elimination efficiencies are lower in surface waters than in STPs owing to shorter residence times, dilute concentrations, and smaller microbial populations. The rate of biodegradation in marine water appears to be lower than in fresh water. APE1&2Cs persist and accumulate. Most fresh and marine water NPE metabolite concentration values are between 0 and 10 ug/ L. The highest fresh water value reported is for NP at 4000 pg/ L (effluent from an anaerobic treatment pond). The NP concentration quickly decreases to 2 ug/L in the Savannah River, less than 10 km downstream from the point of discharge (mainly dilution). APB1&2 that sorb and settle with solid organic material may be transformed to AP under anaerobic conditions in sediment. Partitioning is greater in mud, with a higher fraction of organic matter, than in sand. The lipophilicity and a correspondingly high distribution coefficient of AP favor its persistence and accumulation in sediments that are not disturbed. APE metabolite concentrations in sediment are more widespread, with measurements for NPE0-2 varying from 18 ug/kg (ppb) in a marine lagoon in the Adriatic Sea to 24,350 mg/kg in the Glatt River. The approximate order of elimination efficiency in groundwater, from greatest to least is: APEz, APE1, AP, APEZC, APE1C. The greatest reduction in concentration in the aquifer infiltrated by the Glatt River is observed within the first 2.5 m of the aquifer. Overall elimination efficiencies are greater during the warmer summer season, pointing to the importance of temperature, ranging between 4 and 20°C, in controlling biodegradation rates. 108 Temperature ceases to be an issue at a well located 7 m from the river bed, where a constant temperature between 10 and 12°C is maintained. Though high readings are occasionally seen for NPEz, NPE1, and NP, their concentrations tend to be below 10 [lg/L, and are usually less than 1 pg/ L. Fluctuations in concentration measurements may be due to sudden packets of pollution that overburden the system, or to depletion of dissolved oxygen. Sorption of APEz, APF.1 , and AP retards the movement of these compounds through aquifer material, allowing more time for biotransformation to occur. APE1&2C are the most soluble of the metabolites, and therefore most mobile and least likely to sorb or biodegrade. NPB1,2C concentrations measured in the aquifer underlying the Glatt River are as high as 29 ug/L This concentration is reduced to approximately 5 ug/ L or less within the first 15 m of the river bed. Both biodegradation and dilution contribute to the decrease in APB1&2C concentrations. 109 Halogenated Derivatives One of the final steps in water treatment, particularly in the United States, is disinfection through the addition of chlorine. While destroying waterborne pathogens, the introduction of chlorine and other halogens, that are present with the chlorine as impurities, leads to undesirable reactions with organics that persist in the treated water, or that are present in the receiving waters as humus material. Trihalomethanes - such as CHC13, CHClzBr, CHClBrz and CHBr3 -- are the predominant organic halogens formed, and are the most studied (204-206); chloroform (CHC13), believed to be carcinogenic, is most prevalent. For this reason, studies that investigate the formation of trihalomethanes far outnumber studies that investigate the formation of other classes of organic halogens, such as the halogenated phenols and the halogenated derivatives of alkylphenol ethoxylates. Yet worldwide concerns about the persistence, accumulation, and toxicity of halogenated phenols and halogenated APE metabolites warrant investigations into their formation, behavior and fate. Halogenated phenols and halogenated derivatives of alkylphenol ethoxylates have been found in secondary effluents (1 8, 1 90). Halogenated phenols have been found in the llobregat River, near Barcelona, Spain (207), and halogenated derivatives of APEs have been found in tap water from a water treatment plant in Barcelona (1 9, 20, 22). The Llobregat River is heavily polluted, serving a population of 3.2 million (21), in addition to many industries -- textile, dyeing, tanning, leather, and surfactant manufacturers (21, 22) -- that discharge their effluents into it. Salt mines located upstream are the source of high levels of bromide that are found in the river (21). Chlorination is the means of disinfection in the Barcelona water treatment plant, as well as in the U. 5. plants studied by Reinhard (190) and Ball (18). In 110 Figure 25, Figure 9 has been adjusted to show the effect that chlorination in a water treatment plant has on APE products. A summary of the major findings from halogenated APE research follows. 0 Aerobic transformation of APE metabolites proceeds more quickly and more completely than anaerobic transformation. Under both conditions, APEz and APEZC accumulate as precursors to the halogenated forms (1 8, 23). n OmCHZ-CHZ-o), H p - 4-20 (APEp) aerobic wwt anaerobic wwt R_©—O'(CH2'CH2’O')P-1 CH2000H Fl Omcnzmzo), H p-1-3(APEpC) p-1-3(Ape,,) Fi -O-O—(CH+2-CH2-O-)p H anaerobic wwt/sit p - 0-3 (APEp) \F:naerobicO slt aerobic chlorination @— (AP) R - branched alkyl, M 0(CH2'CH2'O‘)p-1 CHzcOOH usually C3 or C9 p - 1-3 (XAPlEp C) wwt - wastewater treatment IX slt - sludge treatment Lpn -—O-(CH2-CH2-O-)p H x " 0 °' B' p - 0-3 (XAPEp) Figure 25. Proposed APE pathways in an STP with chlorination. Adapted from (15). o Bromide and bromide derivatives are highly reactive. The presence of bromide speeds the rate of transformation to halogenated derivatives (1 8, 1 90, 204). Formation of brominated products is significantly faster than formation of chlorinated products (1 8, 20-22). 111 0 The aromatic structure is preferentially brominated during wastewater chlorination (18, 20, 21, 190). This is depicted in the results summarized in Table 28; concentrations of halogenated OPEps and OPEpCs that were measured in secondary effluents in a water treatment plant in Palo Alto, California (18) are listed. Grab samples were collected three times a day - 9 am, 2pm, and 7 pm - for five consecutive days, and mixed into daily composite samples on a flow-weighted basis. The second column summarizes the results of sample measurements on days 3 and 4, the two days with the highest load. The information in column 4 of Table 28 denotes the number of days for which the compound of interest could be quantified, not quantified, and not detected. Table 28. Concentrations of Halogenated OPEs and OPECs Measured in Secondary Effluents in Palo Alto, California 1 Concentration Concentration Number of days with data; Compound (5 2 day avg) (s 5 day avg) days flat Quantified; (149/ L) (149/ L) days Metected 2 BrOPE1C NQ; NQ M2 0; 5 days NQ BrOPEzC 18 :t 1.4 8.1 :t 9.1 5 BrOPE3C 10.7 :t 6.1 10.7 :1: 6.1 2; 1 day NQ; 2 days nd BrOPE4C nd; NQ nd 1 day NQ; 4 days nd CIOPEZC 5.9 :t 4.0 4.0 :l: 3.1 4; 1 day nd CIOPE3C nd; 6.1 6.1 1; 1 day NQ; 3 days nd CIOPE4C nd; 3.9 3.9 1; 4 days nd BrOPE1 0.9; NQ 0.9 1; 1 day NQ; 3 days nd BrOPEz 1.5; nd 1.5 1; 4 days nd ClOPEz nd; 0.5 0.5 1; 4 days nd 1 Ref. (78). 7- Detection limit 0.5 pg/L. o The biotransformation of halogenated APEp and APEpC (p > 2) leads to the production and accumulation of XAPEZC (X a halogen group), which resists further biodegradation under both aerobic and anaerobic conditions ( 23). 0 Halogenated APs accumulate as a result of anaerobic biodegradation of APE and APEC (23). 112 0 Dechlon'nation does not reverse halogenation. Halogenated derivatives remain in treated water even after dechlorination . 0 High concentrations of XAPEC (60 ppm) are inhibitory to an anaerobic methanogenic consortium (23). Many questions about halogenated organics remain unanswered. The exact mechanism by which halogenation occurs is not known. The factors that influence the extent of halogenation (18) and reaction kinetics (204) are not understood. Toxicity data specific to halogenated APE derivatives were not found. The extent to which these compounds persist, accumulate, and are present in various environmental compartments needs to be investigated. Additional research is required to determine the answers. Toxicity As is the case with determining environmental levels of surfactants, several approaches have evolved to determine toxicity, the factor that is used to satisfy the HAZARD part of the risk assessment equation (RISK a HAZARD * EXPOSURE). Acute toxicity is measured in terms of LC 50 for organisms, or EC50 for plants. Chronic toxicity is measured in terms of no observed effect concentration (NOEC), or lowest observed effect concentration (LOEC). Further, chronic toxicity is determined in relation to some observable, measurable effect, such as growth (length or weight), maturity, spawning, fertilization, reproduction, development, hatch rate, or survival rate. Theoretically the bioconcentration factor (BCF) is the ratio of the concentration of a test compound that partitions to a particular animal or plant organism to the concentration of the test compound that remains dissolved in the water phase. Definitions of these terms are provided in 113 Appendix B. Determinations of chronic toxicity are preferred over acute toxicity when assessing environmental safety (208); values for both are provided in Appendix E (Table 32). Table 29 summarizes the data in Appendix E by providing ranges (from low to high) of estimated toxicity values according to environmental conditions -- fresh water, sediment, or marine water. This facilitates the process of comparing toxicity data to the environmental levels found in Tables 10 and 22 - 24. Bioconcentration factors are provided in Tables 30 and 3 1. Highlights of the findings about AE and APE toxicity follow. 0 The toxicity ranking of linear and branched AEs, APEs, and recalcitrant degradation products (APs), from most to least, is according to: AP > AB (lin) > AB (br) > APE3-20. This relationship is true for both aquatic animals and plants (algae) that have been tested in fresh water and marine water environments (Appendix E). The toxicity of AE and APE nonionic surfactants increases with increasing hydrophobicity and decreasing hydrophilicity. The larger the hydrophobic group, and the shorter the ethoxy chain, the more toxic is the compound (142). linear AEs are more toxic than branched ABS and APEs, probably because cell membranes are impermeable to the bulky branched or aromatic structures. However, linear AEs degrade rapidly and are present in environmental samples at concentrations well below toxicity thresholds. Branched ABS and APEs are an environmental concern because they degrade slowly, and thus are released with their metabolic products to surface waters in secondary effluents (1 7, 142, 181). Branched AEs are used in lesser quantities and are not considered to be a major problem. APs accumulate as biodegradation products of APEs. They are the most toxic and are the greatest concern environmentally. 114 Table 29. Toxicity of Nonionic Surfactants to Aquatic Organisms - Summary (pg/L) Compound Ranges-Acute LCso Ranges-Chronic Ranges-Chronic Environment NOEC LOEC NP Fresh water 791 - 9202 7.49 - 241° 149 - 391° Sediment (pg/kg) 26,1003 - 260,0004 - - - - Marine water 435 - 3006 3.911 - 6.712 6.711 - 5615 AE ' CmEp (1) Fresh water 4602'16 - 29207-15 40013.20 - 20009,"5 100013'1‘1-20 Marine water 7108-17 - 56003'18 - - - - AE ' CmEp (13) Fresh water 61007.19 - 100013-19 - 200013"19 - 115001.19 400014.19 400014.19 NPE9 22 Fresh water 46007 - 14,0001 100013 - 10,0001‘1 200013 - 10,00014 Marine water 12308 - 29.600331 - - - - 1 Water fleas, 48 hrs. 7- Rainbow trout juveniles, 96 hrs. 3 Midge larvae, 14 days. 4 Tadpoles, 10 days. 5 Shrimp, 96 hrs. 5 Shrimp, 96 hrs; this is 2nd highest value for marine water organisms; see Appendix E for further explanation. 7 Fathead minnows, 96 hrs. 3 Shrimp, 48 hrs. 9 Fathead minnows; survival. 1° Water fleas; reproduction rate. 11 Shrimp; length at 28 days. 12 Shrimp; survival and reproduction. 13 Fathead minnows; growth. 1‘1 Water fleas; growth. ‘5 Mussels; growth and strength at 32 days. 15 C1445E7. 17 C13E10Cl. 13 C10E4. 19 C13E7(b). 2° C12-15E9. Z1 NPE90503'. 22 Avg. E0 units/molecule - 9; E0 chain varies between 1 and 20 E0 units/APE molecule. 0 linear AEs: Both acute (1 7, 142, 188, 209) and chronic (1 7, 209) toxicities of primary linear alcohol ethoxylates are eliminated during secondary wastewater treatment. Maki et al. (209) evaluated toxicities in three different solutions - carbon-filtered tap water, stream water, and secondary effluent - and observed that 99% of all mortalities occurred during the first 24 hours, indicating that AE degradation products are not lethal. Further, mortalities were too low to determine LC 50 values for secondary effluents. Their findings corroborate either central fission or (ii-oxidation of the alkyl chain, rather than step-by-step elimination of the ethoxy units, as the probable pathway for biodegradation (1 7). i 115 o Branched ABS and APEs: Kravetz et a1. (1 7) found that chronic toxicity is not eliminated during secondary treatment of branched AEs and APEs. Step-by-step elimination of the ethoxy units is the preferred pathway when the hydrophobe is a highly branched aliphatic or an aromatic associated with a highly branched alkyl chain. This creates products with shorter ethoxy chains and greater hydrophobicity. Not only are these products more likely to sorb to particles that settle with sludge or sediment, but they also degrade more slowly, are more likely to exhibit aquatic toxicity (1 7, 142, 209), and, in the case of APEs, are precursors to APs. o APs in fresh water: Measured concentrations of nonylphenols in fresh water occasionally exceed acute and chronic toxicity thresholds. Acute LC 50 and chronic NOEC and LOEC toxicity values in Table 29 range from 79 to 920, 7.4 to 24, and 14 to 39 (lg/L, respectively. - From the Glatt River study (133), NP concentrations in three out of 1 10 samples were 8 ug/ L or higher, Table 22. Two samples measured at 8 ug/ L, and one sample measured at 45 ug/ L. -- NP concentrations measured in receiving waters of the discharge from a textile mill's waste stream exceed toxicity thresholds in a small creek and in small rivers within 7.2 km of the point of discharge; NP concentrations were 3000, 200, and 30 ug/ L at respective test points 1.6, 2.4, and 7.2 km downstream. The final NP concentration (2 pg/L) in the Savannah River, 9.6 km downstream from the source, was below toxicity thresholds (136). -- NP concentrations in lake Geneva, Switzerland (14), and the thirty rivers study in the United States (54), were below all toxicity thresholds (Table 23). 116 o APs in marine water: From the data available, it is uncertain whether concentrations of nonylphenol derivatives of NPE exceed acute and chronic toxicity thresholds in marine water. Acute LC 50 and chronic N OEC and LOEC toxicity values for marine water range from 43 to 300, 3.9 to 6.7, and 6.7 to 56 ug/L, respectively (Table 29). - Kvestak et a1. measured NPEp concentrations in sewer lines that transport untreated wastewater, at the sewer outlets, and in Sibenik Harbor, Croatia (Table 24). The highest risk of contamination from NP is near the sewer outlets, where dilution, sorption and sedimentation play a role in NPEp removal from the water column. A measurement of the NP concentration at the sewer outlet is not provided. The NP concentration at the sewer outlet following dilution was estimated roughly to be 2.6 ug/L by examining the values provided in Table 24, calculating a ratio of total NPEp concentration at the outlet to NPEp concentration in the sewer line (302/580 - 0.52), and applying the ratio to a sewer line NP concentration value (5 ug/ L) that was obtained from Figure 3c in (137). From this estimation, the concentration of nonylphenols is below toxicity thresholds, but is close enough to suggest a need for additional research. Given the high degree of partitioning to solid fractions and subsequent settling, already discussed, an examination of sediment concentrations of NPE derivatives, especially NP, may be more meaningful. Given the rate at which NPEp is introduced into the harbor, it is likely that NP concentrations are accumulating to toxic levels in the sediments nearest to the sewer outlets. NP concentrations in the Venice Lagoon, Italy (138), were below toxicity thresholds (Table 24). 117 o APs in sediment: Acute Lcso toxicity values for sediment range from 26,100 to 260,000 ug/ kg (Table 29). Determinations of chronic tordcity thresholds in sediments were not found in the literature. -- Environmental levels measured in research conducted in the United States (Table 23) and Italy (Table 24) are well below determined acute toxicity thresholds. - NP levels as high as 13,100 ug/kg have been measured in Glatt River sediments (Table 22). This is below the acute toxicity threshold, but should be monitored. -- The minimum chronic toxicity value for NP in both fresh water and marine water environments in Table 29 is about an order of magnitude less than the minimum acute toxicity value. If this relationship holds for sediments, then NP concentrations in sediments could exceed chronic toxicity thresholds. 0 A higher concentration of suspended solids or colloidal material can decrease the toxicity of a surfactant to some organisms (209) while increasing toxicity to others (25). This is because a surfactant's sorption to suspended solids decreases its bioavailability (209) in the water column, but increases its bioavailability to bottom dwelling organisms that may feed upon particles associated with surfactants (25). 0 Of the organisms tested, the fathead minnow is the most sensitive of the fresh water organisms (1 7, 54), and the shrimp is the most sensitive of the marine water organisms (54) (Table 29). - Patoczka and Pulliam (142) determined the acute toxicity of AEs and APEs in continuous-flow activated sludge reactors to Mysidopsis bahia, a marine organism. They observed that toxicity increases when the E0 chain of a linear AE molecule is capped with a chlorine group (Appendix E). 118 Contrary to this, polarity increases and toxicity decreases when a phosphate or sulfate group is added to the end of the E0 chain of an aromatic based surfactant (Appendix E) . Determination of bioconcentration factors provides another means to evaluate the environmental safety of a compound that appears to be safe for a particular organism based on acute and chronic toxicity tests. It is used to assess potential harm to consumers of that organism, to consumers at higher trophic levels, or long term deleterious effects to the organism itself (210). Ahel et a1. (26) estimated BCFs of fresh water organisms under field conditions. They measured NP, NPE1, and NPEz concentrations in predominant algae, three species of fish, and a wild duck species found in and along Chriesbach Creek, a Glatt River tributary that also receives secondary effluents. Only the most abundant alga, Cladophora glomerata, was sampled from the Glatt River; algal BCFs were estimated for the most polluted Glatt River locations, 15 and 30 km downstream from the river's inlet at Greifensee. Fish and algal BCFs are summarized in Table 30. Concentrations varied widely according to compound, organism, time of year, and location (26). High concentrations of NP and the lower oligomers, NPE1 and NPEz, occurred in the algae examined. Mechanical removal of the algae that covers the river at summer's end removed a substantial amount of surfactant derivatives that were sorbed to the algae; concentrations measured in algal samples collected during autumn were five to twelve times lower than concentrations measured in samples drawn during the summer (data not included). Ahel et a1. attribute this to lower exposure times for the fall samples. The BCFs estimated for algae collected from Criesbach creek were generally lower than BCFs in samples taken from the Glatt River (Table 30). Lower BCFs were estimated for the fishes than the algae (Table 30), indicating that biomagnification due to ingestion of 119 algae was not occurring in these organisms. Concentrations in the duck were comparable to concentrations measured in the fishes (data not included). Although ecological risk cannot be determined based on this information alone, bioconcentration is sufficient to warrant concern about the potential harm of these compounds to man (26). Table 30. Bioconcentration Factors of Nonylphenol Ethoxylate Derivatives in Fresh Water Animals and Plants 1 BCF Species Tissue NP NPE1 NPE; W 2 '15—“ Squalius cephalus Heck. 3 muscle 46 8 14 Squalius cephalus Heck. 4 9““ "“3" 9"“ 118-359 15-78 3-149 Barbus barbus L. 5 muscle 97 135 245 Barbus barbus L. 5' gut, liver, gills 8-251 3-38 3-15 Salmo gairdnen' 5 muscle 38 18 5 Salmo gairdnen' 5 gut 410 304 319 Algae Cladophora glomerata 9750 205 460 Fontinalis antipyretica 1080 40 65 Potamogeton crispus 640 50 200 Glatt River 7 Downstream A13; Location Cladophora glomerata 3 15 km 7700 5000 1800 Cladophora glomerata 3 30 km 6600 3500 1000 1 Ref. (26'). 2 To determine BCFs: Concentrations of NP, NPE1 and NPE; in water from Chriesbach creek are 3.9, 23, and 9.4 ug/L, respectively. 3 Sample size, n - 4. 4 Range of BCFs determined from two composite samples where n - 4 and n - 3; high value is estimated BCF in liver. 5 n - 2; 5' range of BCFs determined for n - 2; high value is estimated BCF in liver. 5 n - 1. 7 To determine BCFs: Concentrations of NP, NPE1 and NPE; in water from the Glatt River at 15 and 30 km are 3.3, 16, and 16 pg/L and 1.1, 4.0. and 4.7 ug/L, respectively. 3 August data. Ekelund et a1. (25) designed laboratory experiments with running seawater to determine bioconcentration factors for 1‘1C labeled 4-nonylphenol in three marine organisms -- Mytilus edulis L. (common mussel), Crangon crangon L. (common shrimp), and Gasterosteus aculeatus L. (stickleback). They measured NP concentration in the total organism, rather than in 120 individual tissues. They conducted two experiments, and for each experiment estimated two bioconcentration factors for each organism, based on tissue weight and fat weight. Their results are provided in Table 3 1. Except for shrimp, their BCF estimates are considerably higher than those provided for fresh water organisms (Table 30). Table 31. Bioconcentration Factors of Nonylphenol in Marine Organisms 1’2 BCF Test 1/T est 2; flanism Compound in tissue (t) or fat (f) Crangon grangon L. labeled NP 1 10/90; (t) shrimp 7500/ 5500; (f) Gasterosteus aculeatus L. labeled NP 1200/ 1 300; (t) stickleback 17,800/16,700; (f) Mytilus edulis L. labeled NP 2740/4120; (t) mussel 169,300/21 6,600; (f) 1 Ref. (25). 2 Test conditions: dosing - 16 days; elimination - 32 days. Beyond concerns about toxicity and bioconcentration of APE metabolites is an ever-growing body of evidence that some alkylphenols, as persistent metabolites of APE surfactants, may contribute to more deleterious ecotoxicological damage. Within the last 20 years, a variety of anthropogenic chemicals that may disrupt biological processes by mimicking or blocking the processes of naturally produced hormones, such as estrogen, have been identified. Several pesticides, polychlorinated biphenyls, combustion pollutants, and plastics ingredients, as well as alkylphenols (penta- to nonylphenol) have been named as contributors to this growing concern (30- 32). Though research related to the estrogenic effects of alkylphenolic compounds is in its infancy, recent findings provide evidence that these compounds are at least weakly estrogenic (28, 33, 34). 121 White et a1. tested 4-octylphenol, 4-nonylphenol, 4-nonylphenoxy- carboxylic acid, and 4-nonylphenol diethoxylate to determine the strength of estrogenicity (33). They found that the order of estrogenicity is: OP> NPE1C> NP-s NPEz. Ethoxylated alkylphenols with more than three ethoxy units did not display estrogenic behavior. An assay test for evaluating estrogenic potency provides evidence that OP is less potent than 17p-estradiol, a natural estrogen; OP stimulates vitellogenin secretion from hepatocytes of male trout to about the same extent as 17p-estradiol when its concentration is 1000 times greater - 10" 5 M for OP vs. 10'8 M for 17p—estradiol (28). NP is between 10 and 20 times less potent than OP. These compounds also stimulate breast cancer cell growth and transcriptional activity in the estrogen receptor; they are mediated by the estrogen receptor and compete with 17p-estradiol for binding to the estrogen receptor (33). Concentrations much higher than those found in the environment are required to initiate a response in in vitro estrogen assay systems. The effect of chronic exposure to these compounds as they accumulate over long periods of time is unknown (34). To date there is no evidence that the endocrine disrupting behavior of the alkylphenolic compounds has caused irreversible damage to any species (34), but their widespread use in large quantities is reason for concern. Concluelone and Recommendatlons Biodegradation and sorption are the major mechanisms that determine the environmental fate of alcohol ethoxylates and alkylphenol poly- ethoxylates. Alcohol ethoxylates are removed readily from wastewaters, biodegrading by central fission or cry-oxidation. Although results are not consistent across all studies, relatively recent data suggest that sorption plays an important role in removing long chain AEs from the aqueous phase. Conflicting results from partitioning studies, the detection of AEs in surface water samples, and higher toxicity levels of parent AB surfactants dictate the need for additional research in this area. The behavior of the alkylphenol polyethoxylates (APE3-20) and their metabolic products (APE1&2, AP, APE1&2C) varies with environmental conditions. The major pathway of biodegradation is step-by-step elimination of the ethoxy units. Widely divergent conclusions regarding the extent of APE elimination, the percentages of APc released to the environment, the effects of sorption, and the extent to which toxic metabolites (AP, APE1&2) accumulate point to the complexity of APE environmental behavior and fate. Though a plethora of data has been collected, knowledge about partitioning behavior, interactions with colloidal material, and the behavior, fate, and environmental effects of carboxylated APEs and halogenated derivatives of APEs is lacking. The distribution, persistence and accumulation of alkylphenol ethoxylate products, especially the octyl- and nonylphenol ethoxylates, is cause for concern in consideration of: APE1&2C concentrations found in river water and AP concentrations found in digested sludge, algae and sediments; the partitioning characteristics of the lipophilic alkylphenol and alkylphenol mono- and diethoxylate derivatives; the likelihood that anaerobic conditions will prevail in sediments, thus favoring biorransformation of APE1&2 to AP 122 123 after settling; the likelihood that resuspension under fast flow conditions will reintroduce metabolites to the water column; evidence that APEl &2C are not readily removed from groundwater; evidence that persistent, toxic halogenated derivatives form during chlorination of secondary effluents and ensuing river transport; evidence that halogenated derivatives are present in drinking water; the possibility that the persistence and accumulation of OP and NP, halogenated and nonhalogenated, could lead to levels that exceed toxicity thresholds determined for algae and aquatic organisms; and evidence that OPE and N PE metabolites may disrupt endocrine behavior. In order to alleviate existing deficiencies in knowledge about the interrelationships of biotransformation and physical-chemical processes of alcohol ethoxylates and metabolites of alkylphenol ethoxylates, future research should focus on the following areas: 1) Further development of analysis techniques for extracting, separating, and identifying nonionic surfactants and their metabolites, including halogenated derivatives, at trace concentrations in complex environmental matrices; 2) Effect of colloidal material on biotransformation, transport, and fate of ABS, and derivative products of APEs, halogenated and nonhalogenated; 3) Partitioning/sorption behavior of ABS, to ascertain that the alcohol ethoxylate parent surfactants are not accumulating to toxic levels in environments with high concentrations of suspended solids or colloids; 4) Determination of the extent to which alkylphenols and carboxylated alkylphenol ethoxylates, halogenated and nonhalogenated, are accumulating in fresh water and marine water sediments; 5) 6) 7) 8) 9) 124 Better definition of the behavior and fate of carboxylated APEs, halogenated and nonhalogenated, through determination of: partitioning behavior in fresh and marine waters; the effect of the concentration of suspended solids and colloidal material; removal efficiency in surface waters and groundwater under warm and cold temperature conditions; extent of accumulation in surface water and groundwater environments; Determination of the toxicity thresholds for carboxylated APEs, halogenated and nonhalogenated; Determination of the fate of APs and the lower oligomers when land applied in digested sludge or algae that has been mechanically removed from natural waters; Determination of the mechanisms in effect when volatilization and photolysis remove APs from surface waters; determination of disposition of APs if volatilization is the removal mechanism; Determination of the esn'ogenic effects of APE metabolites to aquatic organisms and ultimately, to man; 10) Selection of an alternative surfactant for future replacement of the alkylphenol polyethoxylates. APPENDICES APPENDIX A APPENDIX A SEWAGE TREATMENT PLANTS After cleaning is complete the wash water is flushed to a distribution system that transports the water to an industrial or municipal wastewater treatment plant. A discussion about the flow of sewage through a wastewater treatment plant provides an understanding of how surface waters, groundwaters, and soils are also affected. Most sewage treatment facilities are designed to provide secondary as well as primary treatment. Some also offer tertiary treatment. Figure 4 is a generalized flow chart of the activated sludge sewage treatment plant seen commonly in the United States (5 6). Biofilters would replace the aeration tank in a flow chart for a trickling filter plant. An important aspect of the sewage treatment process is sludge handling (dewatering) and disposal. Approximate times for the various waste processing segments are provided in Table 3. Typical suspended solids concentrations are in Table 4. Primary treatment involves the physical removal of solids by screening, skimming, and/ or settling. Screening removes the largest solids, and occurs before the influent enters the primary settling tank. Skirnming removes floating solids; settling removes settleable solids. The purpose of the primary sedimentation tank, or clarifier, is to reduce the load going into the secondary treatment stage. Primary settling can remove up to 7096 and 4096 of the suspended solids and organic matter (as measured by BOD 5), respectively, from influent wastewater. This occurs because low solubility organic compounds are more likely to sorb to the solid organic material that tends to settle in the primary clarifier. The settled material is removed periodically from the bottom of the primary settling tank, thickened, and placed in a 125 126 APPENDIX A digester, where biodegradation continues. Supernatant from the thickened sludge can be cycled back to the primary clarifier for additional treatment. Effluent from the primary clarifier is routed to secondary treatment, which removes organic matter biologically or chemically. Conventionally, secondary treatment is a two step process that involves biological treatment and sedimentation. Suspended growth (e.g., activated sludge), or attached growth (e.g., trickling filter) systems, or a combination of the two, are used for biological removal of organics. The activated sludge process is chosen most frequently worldwide for biological treatment of municipal wastewater. Biological removal occurs in the aeration tank; settling occurs in the secondary clarifier (Figure 4). Mean cell-residence time and food to microorganism ratio are used initially to design the activated sludge system, and later to control the activated sludge process. All of the aeration tank's contents are referred to as the "mixed liquor". The most critical factor in successful biological treatment is maintaining the aeration tank environment to optimize microorganism growth, by monitoring the food-to-microorganism ratio (F/M) and the mixed liquor volatile suspended solids (MLVSS). The microorganisms convert organic waste to C02, NH3, new cells, and other end products. Aeration (by diffusion or mixing) provides enough oxygen and mixing to ensure continued successful operation. Additionally, pH, alkalinity and temperature must be monitored closely. Effluent from the aeration tank includes suspended solids and organic material that are settled in the secondary clarifier. This forms a sludge that is removed periodically. Part of the sludge is recycled to the aeration tank, thus maintaining a constant number of microorganisms, or MLVSS, in the aeration 127 APPENDIX A tank. The remainder is wasted; this portion is thickened before placement in a digester for further stabilization prior to disposal. The supernatant from the thickening process is recycled back to the primary clarifier for additional treatment. The effluent from the secondary clarifier is disinfected prior to release to surface waters to ensure against contamination by pathogenic organisms. Once in the surface waters, components or their metabolites may biodegrade further; or they may sorb to particles and settle to the river or lake bed. Depending on the level of the water table, an underlying aquifer may also be affected (Figure 3). APPENDIX B APPENDIX B GLOSSARY OF TERMS acclimation/ adaptation - process by which an organism is able to revise its metabolic capabilities so as to live and propagate on an organic compound to which it has not been previously exposed (6); advection - the process by which solutes are transported by the bulk motion of the flow of groundwater (5 9); bioaccumulation - uptake of compound by an organism from its environ- ment; dependent on bioavailability, concentration in organism is approximately equal to concentration in environment (21 0); bioconcentration - uptake of compound by an organism across its membranes from its environment; concentration in organism will be greater than concentration in environment (210); bioconcentration factor - the quotient of the concentration of a compound at equilibrium in organism divided by the concentration of compound in surrounding aquatic environment (210, 211); biodegradation - destruction of chemical compounds by the biological action of microorganisms - bacteria, fungi, or yeasts; bacteria are the predominant microorganisms found in the environmental compartments where surfactants and builders are most likely to be found (6, p. 7); biotransformation - change in chemical structure of chemical compounds brought about by microbial action (35); colloids - filterable solids (5 6); microparticles or macromolecules that are small enough (1 nm - 1m) to move primarily by Brownian motion, rather than by settling due to gravitational forces; include humic substances, proteins, viruses, cell fragments, and some bacteria in natural waters; provide another organic micro-environment to which organics can absorb, thus affecting environmental behavior and fate (35); dilution factor - the sum of stream flow and effluent flow divided by effluent flow (49); dispersion - the tendency of the solute to spread as it moves along its path (5 9); Ecso - used to measure toxicity of a test compound to plant life, usually algae; for any given test compound, it is the effect concentration at which the specific growth rate for a culture medium with the test compound is 5096 of the specific growth rate for the culture medium without the test compound ( 21 2); food-to-microorganism ratio (F/ M) - BOD concentration divided by the product of the hydraulic retention time and the concentration of mixed liquor volatile suspended solids (5 6); homolog - refers to variations in the number of carbon atoms in the hydrophobic (alkyl) moiety of a surfactant; hydraulic retention time - as related to activated sludge systems, the volume of the aeration tank divided by the daily influent flow (5 6); 128 129 APPENDIX B inherent biodegradability - a classification assigned to chemicals which may biodegrade, though not within the standardized test periods set for ready biodegradability, indicating need for a period of acclimation (6, 83); the tests: more closely represent environmental conditions, with higher biomass and lower test compound concentrations; allow for the occurrence of cometabolism by providing longer exposure times and multiple carbon sources; measure DOC die away or C02 evolution (aerobically) or COz/CH4 (anaerobically) (83); isomer - refers to structural variations possible in the hydrophobic moiety of a surfactant (i.e., branching configurations, aromatic ring placement on an alkyl chain, or para-, ortho-, or meta- arrangement on an aromatic ring); LC50 - used to measure acute toxicity of a test compound to an organism; the lethal concentration of test compound is the concentration at which 50% of the population of organisms of interest is killed after exposure during a set period of time; tests are set up to run for 24, 48, 96, or more hours, depending on the sensitivity of the organism; LOEC - lowest observed effect concentration; used to measure chronic tordcity; the lowest concentration at which an effect (such as to growth, development, time to maturity, extent of maturity, spawning, fertilization, reproduction, hatch rate, or survival rate) is observed; mean cell-residence time (Be) - as related to activated sludge systems, the mass of organisms in the aeration tank divided by the mass of organisms removed from the entire system per day (5 6); mineralization - the complete conversion of an organic chemical to carbon dioxide, water and other stable inorganic forms of the elements, primarily nitrogen, sulfur, and phosphorus (35, p. 485, 79); mixed liquor volatile suspended solids (MLVSS) - the mass concentration of active rrricroorganisms in an aeration tank (5 6); NOEC - no observed effect concentration; used to measure chronic toxicity; the highest concentration, detected through tests, at which no effect (such as to growth, development, time to maturity, extent of maturity, spawning, fertilization, reproduction, hatch rate, or survival rate) is observed; Oligomer - refers to variations possible in the size of the hydrophilic moiety of a nonionic surfactant (usually relates to the ethoxy units, although propoxy units may also be involved); may also refer to polyethylene glycol molecules that possess between 2 and 8 ethoxy units; ngo- and mer meaning "few parts" (Greek) (213); partition coefficient - a ratio that describes the relative abundance of a compound between two phases, usually the concentration of a compound associated with an organic phase to the concentration of the same compound associated with aqueous phase - e.g., octanol:water (Kow), hexane:water (Khw), sedimentzwater (Km) (35); 130 APPENDIX B primary biodegradation - the alteration in the chemical structure of a molecule by biological action so that characteristic properties (e.g., foaming or surface tension for surfactants) are no longer evident (6, 79); analytical procedures designed to test for primary biodegradability were the original means to determine the acceptability of surfactants based on biodegradability (2, p. 329, 46, p. 375); ready biodegradability - a classification assigned to chemicals which undergo immediate and rapid mineralization according to any of several standardized preliminary screening tests (84, 186); the tests: are 28 days duration; are run at low biomass, yet high test compound concen- trations; are carried out aerobically in aqueous solution under unacclimated conditions; measure dissolved organic carbon (DOC) die away, COz evolution, or oxygen uptake (BOD) (83, 84); recalcitrance - resistance to degradation by microbial action (96); suspended solids - non-filterable solids, usually greater than 1.2 m in diameter; (5 6); subject to gravitational settling; usually separated into volatile and non-volatile fractions during analysis of wastewater and surface waters; ultimate biodegradation - the complete conversion of a molecule to carbon dioxide, water, inorganic salts, and cellular material (6, 79); currently the accepted measurement (46); APPENDIX C APPENDIX C ESTMA'I'ION OF Row USING FRAGMENT CON STANTS The octanol-water partition coefficient (log Kow) included in Table 6 for NPE1-3C was calculated from structural group contributions as indicated below. In Table 6, log Kow and log Khw values are about equal as the number of ethoxy units increases incrementally for the lower oligomers: log Kow for NPE1, NPEz, and NPE3 is 4.17, 4.21, and 4.20, respectively; log Khw for NPE1 and NPEz is 3.4 and 3.3, respectively. From these experimentally obtained values the ethoxy groups appear to have little effect on partitioning. Therefore log Kow - 4.20 was used for the Table 6 calculation, and provides a reasonable approximation for NPE1-3C. NPE1-3 log Kow: 4.20 Approximated in (100). 0.98 Removal of terminal alcohol group as estimated in (35). - 1.11 Addition of carboxylic acid group as estimated in (35). NPE1-3C log Kow: 4.07 More discussion about this estimation process is found in Schwarzenbach (35 ). 131 APPENDIX D APPENDIX D NONIONIC SURFACTANT BIODEGRADATION PATHWAYS A series of laboratory experiments with thin-layer chromatography (TLC), and infrared (IR) and nuclear magnetic resonance (NMR) spectrometry led Patterson et al. to conclude in 1970 that central fission and (xx-oxidation of the hydrophobe operate simultaneously, but as two independent mechanisms, to degrade a variety of commercial surfactant AE molecules (139, 141). On the other hand, Tobin et al., in 1976, (74) and Vashon and Schwab, in 1982, (86) found that hydrolytic cleavage of the ether bond joining alkyl and ethoxy chains (central fission) was most likely to occur in their systems. Tobin et a1. (74) degraded C12-15E9 in shaker flask cultures, field experiments in a lake environment, and a bench scale activated sludge system. They applied a combination of the Wickbold procedure and gas chromatography (GC) to HBr cleavage products to observe rapid loss of the starting surfactant and disappearance of the entire alkyl moiety, respectively. This, together with the identification of ethoxylate material in the lake experiment and low C02 evolution in the shake flask experiment, was the basis for their conclusion that central fission followed by oxidation of the alkyl chain was the mechanism at work. Vashon and Schwab (86) measured the rate and extent of mineralization (via C02 production), in estuarine water, of trace concentrations of both nonionic and anionic 1‘1C-labeled surfactants: *C16E3, C12*E9, *C16E98, and C16*E3S. The 'S' designates presence of a sulfate group on the terminal ethoxy unit. *C16 E3 and *C16E9S were labeled on the a-alkyl C; C12*E9 and C16*E3S were labeled uniformly in the ethoxy chain. Vashon and Schwab observed rapid mineralization of the a-alkyl carbon but very slow 132 133 APPENDIX D mineralization of the ethoxy chain. They likewise concluded that hydrolytic central fission initiates biodegradation in the AEs. In 1985, Steber and Wierich examined alkyl or ethoxy radiolabeled metabolites (”Q from the biodegradation of stearyl alcohol ethoxylates (C13E7) in a model continuous flow sewage treatment plant. Their results support Patterson's conclusions that central fission and hydrophobic m- oxidation occur simultaneously to degrade AE molecules (140) . Steber and Wierich, 1987, took their work a step further when they ran similar experiments (1 10) in an anaerobic sludge digester. They observed that only the hydrolytic pathway was effected anaerobically (1 1 0), and that hydrolysis occurred either directly or oxidatively. Swisher (6) suggests several alternative oxidative hydrolytic pathways, which are presented in Figure 7 and discussed in more detail in the narrative. Related to the third possible point of attack, Steber and Wierich, 1985, (140) concluded that the presence of dicarboxylated alkyl-14C polygycols and monocarboxylated alkyl-labeled pEO intermediates provides evidence for microbial degradation of the ethoxy chain, although at a much slower, perhaps even time-delayed, rate. Steber and Wierich are joined by others (6, 113, 11 7, 118) in suggesting that degradation of the ethoxy chain occurs in a stepwise fashion, beginning with the terminal monomer unit, similarly to the degradation of polyethylene glycols (PEG's). More details are provided in Figure 8 and the narrative. A pathway that will not be discussed in detail, but deserving mention, is one reviewed by Swisher (6). Swisher cites the works of Zahn, 1980, (214) who deduced carboxylation of the terminal alcohol group of the E0 chain followed 134 APPENDIX D by two one carbon scissions (first decarboxylation, then release of formate), and Schbberl, 1983, (215) whose results support Zahn's deductions. APPENDIX E APPENDIX E TOXICITY OF NONIONIC SURFACT ANTS TO AQUATIC ORGANISMS Table 32. Toxicity of Nonionic Surfactants to Aquatic Organisms Measure of toxicity (ug/ L) 1 A: Acute - LCso or ECso (alga); duration N: NOEC; chronic effect L: LOEC; chronic effect Organism Compound Test conditions Ref. fresh—m Pimephales €12-15E9 A: 1600; 96 hrs. 25°C/intact surf't in (17) promelas (linear) N: 400; growth effluent from fathead minnow L: 1000; growth continuous activ'd sludge biotreater/ EPA procedures C13E7 A: 6100; 96 hrs. same as above (17) (branched) N: 1000; growth , L: 2000; growth €14.15E7 A2: 1130 - 1260; 96 hrs. 20°C/carbon-filtered (209) N: 800; survival tap water/ static toxicity C14-15E7 A2: 1290 ' 1500; 96 hrs. 20°C/ (209) N: 800; survival stream water/ static toxicity C1445E7 AZ: 2090 - 2920; 96 hrs. 20°C/ 2° effluent/ (209) N: 2000; survival static toxicity NPE9(b) A: 4600; 96 hrs. 25°C/intact surf't in (17) N: 1000; growth effluent from L: 2000; growth continuous activ'd sludge biotreater/ EPA procedures NP A: 300 8: 135; 96 hrs. EPA Good Laboratory (54) N: 23; hatch rate Practices protocols L: >23; hatch rate N: 23; length, 28 days L: >23; length, 28 days N: 7.4; survival, 33 days L: 14, survival, 33 days Salmo gairdnen' C1445E7 A2: 460 - 1210; 96 hrs.3 15°C/in effluents (188) rainbow trout- Cur-15511 AZ: 970 - 1240; 96 hrs.3 from biofilters/ juveniles continuous flow NP A: 230; 96 hrs EPA Good Laboratory (54) A: 480; 24 hrs Practices protocols rainbow trout - NP A: 920 & 560; 96 hrs. 15-16°C/static (202) juveniles toxicity in fresh 135 water aquarium 136 APPENDIX E Table 32. (cont'd) rainbow trout - NP AZ: 364 - 604; 24 hrs. 9.5-11°C during (202) embryos exposure; then 9.5 - 28.5‘C/fresh water aquarium/ exposure followed by 48 hrs in clean water Daphnia magna €12-15E9 A: 1300; 48 hrs. 25‘C/intact surf't in (17) water flea (linear) N: 1000; growth effluent from L: 1000; growth continuous activ'd sludge biotreater/ EPA procedures C13E7 A: 11,600; 48 hrs same as above (17) (branched) N: 4000; growth L: 4000; growth NPE9 A: 14,000; 48 hrs same as above (17') (branched) N: 10,000; growth L: 10,000; growth NP A: 190 8: 440; 48 hrs EPA Good Laboratory (54) N: 24; reproduction rate Practices protocols L: 39; reproduction rate Daphnia pulex NP A2: 79 - 201; 48 hrs. 23-27’C/fresh (202) water flea A2: 114 - 238; 48 hrs. water/ incubator Selenastrum €12-14E4 A: 2000-4000; 48 hrs. ECso criteria: (212) capn'comutum C124 4E9 A: 4000-8000; 48 hrs. retardation of green alga (312-14513 A: 10,000; 48 hrs. specific growth rate NPEBJ A: 20.000; 43 hrs. to 50% of that in NPE 8.9 A: 50,000; 48 hrs. culture medium NP A: 410; 96 hrs. EPA Good Laboratory (54) Practices protocols Nitzschia C12-14E9 A: 5-10,000; 48 hrs. ECso criteria: (212) fonticola retardation of blue-green 3193 specific growth rate to 50% of that in culture medium Microcystis C1244E9 A: 10-50,000; 72 hrs. same as above (212) aerugmosa diatom Fresh water gdiment Rana catesbiana NP A: 260,000 ppb (pg/kg); EPA Good Laboratory (54) tadpole 10 days Practices protocols Chironomus NP Sub-A: 26,100 ppb (ug/kg); MATC - Maximum (54) tentans 14 days acceptable toxicant midge larva concentration/ EPA Good Laboratory Practices protocols 137 APPENDIX B Table 32. (cont'd) Maw Mysidopsis CrgEroCl A: 710; 48 hrs. 23-25'C/continuous (142) bahla C13E9.75 A: 2200; 48 hrs. flow act'd 3| ICII’S/ shrimp (11054 A: 5600; 48 hrs. static-renewal / N. J. NPE9 A: 1230; 48 hrs. Admin. Code. 7313 "£90503- A: 29,600; 48 hrs. (amended 7-84) “ma-2 A: 4600: 48 hrs. NP A: 43; 96 hrs. EPA Good Laboratory (54) N: 3.9; length 9 28 days Practices protocols N: 6.7; reproduction N: 6.7; survival L: 6.7; length 9 28 days L: 9.1; reproduction L: 9.1; survival Crangon NP A: 300; 96 hrs. 10‘C/static/HPLC (24) septemspinosa shrimp Salmo salar NP A: 190; 96 hrs. 10°C/static/HPLC (24) Atlantic salmon A: 160 8: 130; 96 hrs. 10°C/flow-through/ HPLC Mytilus edulis NP A: 26004; 96 hrs. EPA Good Laboratory (54) mussel L: 56; growth and strength Practices protocols @ 32 days Skelatonema NP A: 27; 96 hrs. 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