Environmental aspects of the use of NTA as a detergent builder

Water Res. Vo[. 18, No. 3. pp. 255-276. 1984 Printed in Great Britain. All rights reserved

0043-1354 84 53.00 +0.00 Copyright ~ 1984 Pergamon Press Ltd

REVIEW PAPER E N V I R O N M E N T A L ASPECTS OF THE USE OF NTA AS A D E T E R G E N T B U I L D E R R. PERRY, P. W. W. KIRK, T. STEPHENSON and J. N. LESTER Public Health Engineering Laboratory, Civil Engineering Department, Imperial College, London SW7 2BU, England

(RecewedJune 1983)

A~trsct---The behaviour, fate and significance of the detergent builder nitrilotriacetic acid (NTA) has been reviewed with particular reference to the removal of NTA during wastewater treatment and the effects of N T A on heavy metal solubility both during treatment and in the receiving environment. It is concluded that NTA removal during secondary biological treatment is subject to considerable variation, both temporal and between works as a result of changes in NTA load, temperature, water hardness and treatment process parameters. As a result of such variability, effluent N T A concentrations may be sufficient to mobilise heavy metals resulting in metal contamination of receiving waters and potable waters. particularly in areas of low effluent dilution and high water re-use. Removal of NTA during primary sedimentation and septic tank treatment is concluded to be predominantly due to adsorption to the sludge solids while removal in anaerobic sludge digestion is subject to operational characteristics of the treatment works. Disposal of N T A contaminated sludge to land may contaminate groundwaters and affect heavy metal speciation, while the disposal of contaminated sludge or sewage to sea may result in toxic algal blooms, in addition to effects on metal speciation. Key words--nitrilotriacetic acid, detergent builders, water, wastewater, heavy metals, heavy metal mobilisation, biodegradation, wastewater treatment, sludge digestion, sludge disposal

INTRODUCTION Eutrophication and associated problems of deteriorating water quality have received considerable attention in recent years particularly in respect of the effects on freshwater lakes and impoundments which tend to be phosphorus limited (Lee et al., 1978). Measures to control eutrophication have largely been directed at point source emissions such as municipal sewage effluents, although other significant sources of phosphorus entering freshwaters include drainage from agricultural land, excreta from livestock, industrial effluents, atmospheric deposition and diffuse urban drainage (Lee et al., 1978). Indeed the contribution of municipal effluent represented only 21% of the total phosphorus load on the Great Lakes in 1976 (Wendt, 1982), 20% of the total load on Lake Mjosa in Norway prior to reductions in phosphate in municipal effluent (Baalsrud, 1982), and 20% of the total input to the eutrophic coastal Adriatic Sea from the Emilia Romagna region of Italy (Haddrill et al., 1983). Although it has been concluded that a reduction in phosphorus concentration in municipal effluent will not, in isolation, lead to a significant improvement in water quality (Lee et al., 1978) the United States and Canada via the 1978 Great Lakes Water Quality Agreement stipulated that the effluents from 400 municipal treatment plants discharging to the lakes should contain a maximum of 1.0 mg 1- t of phosphorus in the upper lakes and 0.5 mg 1-' phosphorus in the lower lakes, therefore requiring the installation of chemical treatment facilities

(Harrington-Hughes, 1978). In order to reduce the phosphorus content of municipal sewage, voluntary and statutory restrictions have been introduced to limit the use of the detergent builder sodium tripolyphosphate (STPP) (Taylor, 1980). The United States currently has "zero" phosphate legislation for detergents in five states bordering the Great Lakes, limiting the phosphorus in laundry detergents to less than 0.5~o by weight, although the State of Wisconsin in late March 1982 decided not to continue its ban on phosphate detergents, first enacted in 1978, since the Wisconsin Department of Natural Resources was unable to demonstrate an improvement in water quality (Flynn, 1982). In regions which have no detergent phosphate ban in effect only 35°/o of the phosphorus in domestic wastewaters is attributable to phosphate-built detergents (Lee et al., 1978) and a detailed study in Michigan indicated that only 2°,; of the total phosphorus loading on the Great Lakes adjacent to this state could be attributed to laundry detergents if the Michigan detergent phosphate ban were not in effect (Wendt, 1982). In Canada a mandatory 2.2% detergent phosphorus content, resulting in an 80% reduction in phosphates in laundry detergents in 1973, led to a 30% reduction in raw sewage phosphorus concentration in Ontario (Appleby, 1977). The contribution of STPP to sewage phosphorus in the U.K. is estimated to be between 40 and 50% (Department of the Environment, 1978) while in Germany the phosphorus content of detergents was reduced by 22-28% to approx. 7% phosphorus (28o/0

255 W R 1~,3~A

256

R. PERaY et al.

STPP) on 1 October 1981 and will be further reduced minosilicate, softens water by ion exchange and this is generally slower than chelation (Henning et at,, by about 20°0 to around 5.5°~ phosphorus 1977). The insolubility and ion exchange properties (22°0 STPP) by 1 January 1984 ( H a m m e t al., 1982). Prior to these reductions detergent phosphorus con- of zeolite have led to an investigation of the fate of stituted 40°'0 of the total phosphorus content of zeolite during sewage treatment and the influence of German domestic sewage (Hauptausschuss Phos- zeolite on heavy metal removal during sewage treatphate und Wasser der Fachgruppe Wasserchemie in ment. The removal of zeolite during primary sedimentation has been studied (Carrondo et aL, 198 la) der Gesellschaft Deutscher Chemie, 1978). In the early t970s when the reduction or elimi- and was demonstrated to be between 66 and 76°0 nation of detergent phosphorus was advocated in the depending on flow conditions, with slightly improved United States Great Lakes region, a suitable replace- removals of lead, zinc, copper and chromium under ment with detergent performance and safety charac- certain conditiors (Rossin et al., 1982b). No effects teristics similar to those of STPP was not available on the activated sludge process performance or on metal removal in that process have been demon(Rutkowski, 1981). An intensive search indicated that the principal organic contenders were nitrilotriacetic strated (Carrondo et al., 1980, 1981b). Of the suggested detergent builder replacements acid (NTA) as its trisodium salt (Na~NTA), sodium citrate, sodium carboxymethyltartronate (CMT or N T A has received the greatest attention. It was used Builder M) and sodium carboxymethyloxysuccinate to a limited extent in the United States in the late (CMOS) (Rutkowski, 1981), while the inorganic 1960s but in December 1970, at the request of the builder zeolite type A has been adopted by some U.S. Surgeon General, detergent manufacturers voldetergent manufacturers as a partial substitute for untarily agreed to desist from the use of NTA in STPP (Berth, 1978). The International Joint Com- laundry detergents until further health studies were mission (1980) concluded that the use of citrate as a completed. The request was based on a report from detergent builder would not be detrimental to the the National Institute of Environmental Health environment, and indeed was more acceptable than Sciences which indicated that NTA complexes of cerCMT or CMOS. However, citrate is a much less tain heavy metals such as cadmium and mercury may effective sequestrant for hardness ions than STPP, produce mutagenic or teratogenic responses m pregand was not marketed in powdered laundry deter- nant mice (Hammond, 1971). The United States gents up to 1981 (Rutkowski, 198l). CMOS is also government's position against NTA was reaffirmed in readily bidegradable although it requires longer peri- September 1971, and again in May 1972, and once ods for the establishment of suitable microbial popu- more in November 1974. However, in May 1980 the lations and is consequently more susceptible to vari- Office of Toxic Substances decided not to regulate the ations in environmental conditions than is citrate use of NTA in the United States with various pro(International Joint Commission, 1980). Viccaro and visos including a recommendation that it should not Ambye (1977) reported the degradation of CMOS in be used in shampoos, foods, hand dishwashing delaboratory anaerobic digesters after an accli- tergents or similar products, that work places should matisation period of 6 days, removal being virtually be strictly controlled and the environment monitored complete after 14 days. The International Joint Com- if this product should be reintroduced into detergents mission (1980) concluded from a review of the litera- in the U.S. (Birkner, 1982). In November 1982 the ture on CMOS that the introduction of CMOS on a Environmental Conservation Commissioner of New large scale should be preceded by further study, York State announced his intention to prohibit the particularly into the problems of acclimatisation in use of NTA in household cleaning products on the wastewater treatment and heavy metal chelation. basis of possible adverse effects on human health via CMOS is a weak calcium sequestrant compared to drinking water contamination (State of New York, STPP, although slightly superior to citrate and could 1982). Despite the United States' position with regard to only be a partial substitute (Lamberti, 1977) making its introduction unlikely. Barth et al. (t979) found the use of NTA in detergents NTA constituted on that an acclimatisation period of 14 weeks was re- average 15Vo by weight of detergents in Canada in quired before CMT was degraded in the activated 1981 (Perry, 1981). The use of NTA in Canada has sludge process, acclimatisation time increasing at low been estimated to be 55 million pounds in t982. If temperatures (International Joint Commission, NTA was to be generally adopted in the United 1980). Furthermore CMT has not been shown to be States at similar levels to those in Canada it could biodegradable under anaerobic conditions, its use as have a total usage of 627-666 million pounds per a detergent builder was not endorsed by the Inter- annum in the United States (Environmental Protecnational Joint Commission (1980) and its intro- tion Agency, 1980). Finland, Sweden, Switzerland duction appears improbable. Zeolite type A, an alu- and the Netherlands use NTA and the latter, together with Canada and Norway have set drinking water standards at 50 t~g 1-t (50 ppb)* (Environmental Pro*NTA concentrations are expressed as the anion (NTA3-) tection Agency, 1980; Perry, 1981). According to Perry (1981) NTA constitutes less than 1~,i of the except when ori~nally unspecified.

NTA as a detergent builder total weight of detergents in Sweden, Holland and Switzerland. The West German Federal Health Agency has recommended that NTA concentrations in drinking water should not exceed the detection limit of 0.2 ,ug 1-~ (0.2 ppb) (Bundesgesundheitsamt, 1982). NTA has been withdrawn from the Japanese market by the manufacturers (Environmental Protection Agency, 1980). In the United Kingdom it is considered that the curtailment of the use of phosphate in detergents is unnecessary, since "At present there is not a completely acceptable substitute for sodium tripolyphosphate nor is one necessary because sodium tripolyphosphate is completely satisfactory from both the washing efficiency and environmental aspects under United Kingdom conditions and any potential substitute would need to be extensively tested under such conditions", (Department of the Environment, 1980).

NITRILOTRIACETIC

ACID

In terms of washing performance NTA can largely replace STPP and as such is the principal alternative (Perry, 1981). However, NTA has received considerable attention primarily due to its demonstrated carcinogenicity (National Cancer Institute, 1977; Rail and Goyer, 1980) and heavy metal chelating properties (Mottola, 1974; Stoveland et al., 1979a). Reduction in NTA concentration as a result of municipal wastewater treatment has been concluded to be substantially dependent upon secondary biological treatment, and the aerobic biodegradability of NTA has been extensively studied and reviewed (International Joint Commission, 1978). Toxicology

Nitrilotriacetic acid and its trisodium salt have been reported to be carcinogenic to the urinary tracts of both rats and mice (National Cancer Institute, 1977) and tumorigenic to the rat kidney (Goyer et al., 1981). It was demonstrated to have a weak chromosome breaking effect (Kihlman and Sturelid, 1970) and also produced a positive result in a cytogenic assay using Vicia sp. during the U.S. Environmental Protection Agency Gene-Tox Program (Constantin and Owens, 1982). There is some evidence of increased mutagenicity of insoluble compounds of NTA (Loprieno et al., 1982, 1983). Following an appraisal of the literature concerning N T A toxicology the Executive Committee of the National Toxicology Program in the United States concluded that the mechanism of N T A carcinogenesis and the level of exposure that constitutes a threshold, if there is a threshold, are not known (Rail and Goyer, 1980). Furthermore the possibility that NTA may have a cocarcinogenic effect, through its irritation of urinary epithelium, could not be excluded, suggesting that people with intercurrent disease of the urinary tract

25-

or those exposed to other chemicals that reach the urinary tract may be at higher risk to the toxic and carcinogenic effects of NTA (Rail and Goyer. 1980). Acute and chronic toxicity studies of NTA have been undertaken on both terrestrial and aquatic organisms, and the results have been widely reviewed (Thom, 1971; Epstein, 1972; International Joint Commission, 1978; Rubin and Martell, 1980). Since most bioassays, including those with aquatic organisms, are performed with a single test species in a fixed environment, the results of such studies are not easily extrapolated to natural ecosystems (International Joint Commission, 1978). Additionally few human toxicity studies have been reported (Mottola. 1974). As a result of such uncertainties the U.S. Environmental Protection Agency urged "'manufacturers and processors of NTA to limit occupational exposure", and also recommended that, "NTA not be used in consumer products to which there is direct dermal or oral exposure" (Federal Register, 1980). Meiners et al. (1979) reported that NTA could not be detected in the urine of production workers. However the analytical method used unfortunately had a detection limit of 0.5 mg 1-~. M e t a l chelating properties

The strong chelating capacity of NTA might be expected to have adverse effects upon heavy metal removal in sewage treatment processes and to mobilise metals from sediments in receiving waters. Mobilisation in simple systems may be calculated from equilibria data but such methods are not presently applicable to complex natural systems and mobilisation must be determined by observation or experimentation (Allen and Unger, 1980). Cheng et al. (1975) demonstrated a reduction in heavy metal uptake by activated sludge in the presence of NTA. Heavy metal removal during conventional two stage primary sedimentation and activated sludge treatment is tisually 70% or greater (Brown and Lester, 1979; Lester et al. 1979; Stoveland et al., 1979d: Rossin et al., 1983b). Reported increases in the solubilities of cadmium, chromium, copper, iron, nickel, lead and zinc in the presence of NTA are listed in Table I. The concentration of NTA has been identified as the single most significant parameter determining the extent of metal mobilisation during such experiments (Allen and Boonlayangoor, 1978), although pH may also be important (Salomons and Pagee, 1981). In recent studies NTA has been demonstrated to reduce the soil adsorption of cadmium (Elliott and Denneny, 1982) but increase the adsorption of copper(II) (Elliott and Huang, 1979). Granrli and Edler (1983) demonstrated the ability of N T A to chelate copper(II) in coastal seawater thereby reducing toxic inhibition by copper(II) and stimulating the growth of the toxic red tide dinoflagellate Prorocentrum minimum.

R. PERRY ef c1[.

25",

Table ! Mobilisatton of metals from sediments and sev, age b) NTA NTA ling t " I

Reference Bj6rnda[ et al. (t972) Chau and Shiomi tt972~ Gregor !19721 Zitko and Carson (19"2~ Banat et M. {1972,} Allen and Boonla)angoor 1[9781 Stoveland et ul. (1979cl Salomons and Pagee ([gstl Dietz (1982l Rossin et ~d. 11982a~

5.85 t 10 1.46 146 I-lO0 I t00 0.75 It') 0.2 5 1 15

Metals mobilised Cd, Cu. Ni Cu, Fe. Ni. Pb Cu, Fe. Zn Cd, Cu. Ni: Cd, Cu. Fe, Cd, Cr, Pb. Cd. Cu. Ni. CU, Ni. Zn Ni. Pb. Zn

Matrix Serried se,aage Lake sediment Lake Reser,,oir ~edimenV,

Zn

River sediment River sediment River sediment Mixed liquor River sediment River sediment Mixed primary sludge

Pb. Zn Ni. Pb. Zn Zn Pb, Zn

this figure has probably diminished slightly over the past two years. Canadian monitoring experience has been advocated as representative of sewage influent and effluent concentrations. The results of a survey undertaken in Canada from January to December 1972 and from April 1973 to March 1975 have been reported by Woodiwiss et al. (1979). Between these two periods the NTA content of detergents was increased from 6 to 15"o by weight. During the first period the geometric mean influent concentration to 13 Ontario sewage treatment plants was 1.27 mg 1' NTA with a range of <0.018-8.05 mg i-'. Over the second period the geometric mean influent concentration had increased to 2.65 mg 1-~ with a range of 0.008-19.67 mg 1~ . As these values exhibit fair proportionality it appears likely that 3051; NTA in detergents, the greatest amount likely to be present, would give rise to a geometric mean concentration of about 4 . 4 m g l -t and a maximum concentration of about 30 rag l-' (International Joint Commission, 1978). The results of Canadian monitoring exercises are not however directly comparable to the European situation. Water consumption is much higher in Canada on a per capita basis when compared to many other countries, and sewage is consequently much more dilute (Stoveland et al., 1980). Water consumption in Canada at 614 1 per capita day- ~ (135 Imperial gallons per capita day ~) may be compared

INCIDENCE OF NITRILOTRIACETIC ACID IN SEWAGE AND REMOVAL BY NON-BIOLOGICAL TREATMENT

The extent of NTA removal from municipal wastewaters is dependent on the particular sewage treatment process or combination of processes employed. NTA removal during primary sedimentation will determine NTA loading on secondary treatment processes where secondary treatment is practised (Tabte 2). In Europe and North America significant quantities of sewage are discharged directly to the environment (Table 2) or receive treatment on-site by septic systems. Where anaerobic sludge digestion is undertaken the association of NTA with sludge solids during primary sedimentation will determine the NTA loading on sludge treatment. Secondary biological treatment is not universal and treatment solely by sedimentation-chemical precipitation accounts for 51 and 26% of all treatment in Norway and Finland respectively (Melkersson and Stendahl, 1982). [nfluent concentration

The concentration of NTA likely to be encountered in raw sewage has been widely reported in the literature with quoted figures derived from measurements taken in areas where NTA is i,n use and calculations based on estimated wastewater flows and detergent usage. NTA constituted 15~o on average in Canadian detergents in 1981 (Perry, 1981) although

Table 2. Comparison of municipal sewage treatment facilities ?'o Population

Reference Bunch (1982) White (1982) Schmidtke (1982) Balmer (1982)

Country U.S.A U.K.

served by municipal wastcv, ater treatment

°o Treatment plants .

.

.

.

Primary sedimentation alone

Canada

71 95 49{

23 13 5

Sv, eden

83

< I

14~ 22' 60-65

67

Beccari (1982) Italy Van Haute and Brabander (1982) Belgium

.

.

.

.

.

75 77* 33 secondary 48 lagoons 19

.

.

.

.

.

. . . Tertiar) or

.

advanced 2 10f > 6 Biological + chemicai 74 Biological + chemical 5 Chemical

33

Fteekseder (1982)

Austria

Melkersson and Stendahl (t982)

Finland

-

I

13

Melkersson and Stendahl (t982)

Norwa)

-

7

8

*Approx. 50% activated sludge. 50,";; trickling filter. *Lagoons, land irrigation, micros*raining, sand-fihration. {87~q, of population surveyed. :~1971 figure. f Biological.

.

Primary and biological

50 52 Biological + chemical 26 Chemical 5[ Chemical 34 Biological + chemical

NTA as a detergent builder with consumption in England and Wales at 273 1 per capita day -t (60 Imperial gallons per capita day -*) (Van der Leeden. 1975). Using data presented in the 12th progress report of the U.K. Standing Technical Committee on Synthetic Detergents (Department of the Environment, 1971), which indicate an average phosphate content in the influent to the Slough Sewage Treatment Works of 13.8 mg 1-* (as phosphorus), Stoveland et al. (1980) estimated that if NTA were used in detergents at the Canadian level of 15% its concentration in the raw sewage would be 1 3 - 1 4 m g l -*. Higher concentrations would be expected in hard water areas such as parts of Germany, where the phosphate concentration in German domestic sewage is 10-40 mgl -t (as phosphorus) (Hauptausschuss Phosphate und Wasser der Fachgruppe Wasserchemie in der Gesellschaft Deutscher Chemie, 1978) and in areas which separate surface drainage from domestic and industrial sewage. Variations in influent concentration

Variations in the concentration of NTA in raw sewage may be expected to correlate with current concentrations of condensed phosphates which are almost exclusively derived from detergents. Rossin et al. (1982a) reported peak concentrations of condensed phosphates 3 to 4 times the average influent concentration, and associated these peak concentrations with "wash day" discharges. Since such variations would occur with any alternative detergent builder, it was considered probable that with an average influent NTA concentration of 15mgl -t concentrations would vary between 6 and 34 mg 1-t, while an average influent concentration of 30 mg 1-~ of NTA would have a corresponding variation between 10 and 66 mg 1-t (Rossin et al., 1982a). Wei et aL (1979) also reported diurnal variations of NTA in raw sewage at a Canadian waste water treatment plant which averaged 2.5 mgl -~ with peak concentrations of approx. 12 mg 1-* in the afternoon and minimum values of 0.1 mgl -~ at night. In one instance the raw sewage NTA concentration rose from nearly zero to greater than 14 mg 1-* in only 2 h. Removal during primary sedimentation

The removal of NTA during the primary sedimentation of municipal sewage has been the subject of limited study. Thompson and Duthie (1968) reported that NTA had no effect on solids removal. Shumate et aL (1970) observed an apparent NTA removal of

259

8%, although this was attributed to faulty sampling and it was concluded that removal was negligible. Thayer and Kensler (1973) reported 6 month average removals of 30, 59 and 11% for 3 Canadian sewage treatment works where the influent concentrations were 3.50, 1.43 and 1.21 mg I-* of NTA respectively. Unfortunately the sampling regime for this study was not detailed and variations observed in influent and effluent N T A concentration may have been reduced by lengthy composite or unrepresentative discrete sampling. The most comprehensive study of NTA removal during primary sedimentation has been carried out by Rossin et aL (1982a) using two pilot plants operating in parallel to treat raw sewage. Sampling was undertaken every 3 h and influent raw sewage and settled sewage samples were each composited over 3 days for each experiment under various conditions of hydraulic load. A summary of the results is presented in Table 3. No correlation was evident between the removal of NTA and suspended solids removal or Chemical Oxygen Demand (COD) removal, although adsorption onto the sludge solids was cited as the most probable removal process. A reduction in metal removal was observed during the study with a 10-20% reduction in the removal of zinc and a 5-10% reduction in the removals of nickel and lead (Rossin et al., 1982a). Melkersson and Stendahl (1982) demonstrated that the presence of NTA during sewage treatment by chemical precipitation significantly increased effluent turbidity and phosphorus concentrations while allowing more of the precipitating agent to pass into the effluent. Countries such as Norway and Finland, where one and two stage direct precipitation is the predominant sewage treatment process, may thus experience a substantial deterioration in their sewage treatment due to the presence of NTA (Melkersson and Stendahl, 1982). Adsorption of NTA onto sludge solids during primary sedimentation would inevitably increase the load of NTA on sludge treatment and disposal systems. However, the degree of removal would appear insufficient to markedly reduce NTA loading on secondary biological sewage treatment processes. B E H A V I O U R OF N I T R I L O T R I A C E T I C ACID DURING BIOLOGICAL WASTEWATER TREATMENT

"'Reductions in NTA concentrations which occur

Table 3. Removal of N T A during primary sedimentation (Rossin et aL. 1982a) Parameter

System

N T A effectively added (rag I -*) N T A removal (%) Suspended solids removal (";) C O D removal (%,)

Test Test Control Test Control Test

1 DWF* 15.5 26 37 41 22 21

3 DWF 17.0 24 59 59 48 51

*Dry Weather Flow--hydraulic retention time 2.28 h. tVariable flow simulated diurnal variation in flow to the works.

Variable flowt 18.5 16 4,4 36 ---

I DWF

3 DWF

33.5 40 48 58 43 50

34.5 18 64 62 49 41

Variable flow 37.5 29 51 54 ---

2~,1

R. PER.RYet al.

as wastewater flows through treatment plants are ordinarily assumed to be attributed to biological degradation" (International Joint Commission, [978). Thus the removal of NTA is substantially dependent upon secondary' biological treatment. Bouveng et al. (1968) observed 25 and 36°,,0 adsorption of NTA by activated sludge mixed liquor in jar tests when NTA was added at concentrations of 2 and 10 mg l -~ respectively. Rossin et al. (1982c) in batch and dynamic tests found adsorption by unacclimatised mixed liquor to vary between 13 and 4500 . Other workers have observed only minimal adsorption of NTA by mixed liquor solids (Swisher et al.. 1967: Gudernatsch, 1970; Gledhilt. 1977). Regardless of the degree of adsorption this mechanism will be unimportant in the removal once the biodegradation of NTA has commenced as it is only an intermediate stage in the passage of NTA into the bacterial cell, the catabolism of a small molecule such as NTA presumably being an intracellular process (Rossin et al., 1982c). Biodegradability tests on NTA have been inconsistent: 900/0 degradation was observed after 9 and 13 days in activated sludge and "Gledhill" tests respectively while degradation amounted to only 20~/oin a CO,_ evolution test after 28 days and did not occur in shake flask and BOD tests (Means and Anderson, 1981). Following a period of acclimatisation, near complete biodegradation can be achieved under optimum conditions in the activated sludge process (Gudernatsch, 1970: International Joint Commission, 1978; Stoveland et al., 1979a: Rossin et al., 1982c). In laboratory, pilot and full scale studies of the activated sludge process the acclimatisation period has ranged between 6 and 90 days (Table 4). Cleasby et al. (1974) reported an acclimatisation period of 14-28 days for a pilot scale trickling filter. The efficiency of NTA removal during biological sewage treatment and the period of acclimatisation prior to NTA biodegradation are affected by such factors as concentration of heavy metals (Warren, 1974; Shannon et al., 1978; Rossin et al., 1982c, 1983a). treatment temperature (Eden et al., 1972; Obeng et al., 1982); NTA loading (Stoveland et al., 1979a, b; Wei et al., 1979; Obeng et al., 198t; Rossin et al., 1982c, 1983a), water hardness (Bj6rndal et al., 1972; Stoveland et al., 1979a) and process parameters (Obeng et al., 198l: Rossin et al., 1982c, 1983a).

Interaction o f N T A

with heavy m e t a l s

The biological stability of NTA appears to be strongly related to the specific metal-NTA complex as suggested by Warren (1974). Bj6rndal et al. l1972) reported that the dosing of nickel to laboratory activated sludge units resulted in a significant reduction in NTA biodegradation, although copper and cadmium had little effect. In contrast, Huber and Popp (1972) found that equimo[ar additions of cadmium to laboratory activated sludge units receiving 30mg I -~ of NTA resulted in the cessation of NTA biodegradation which only resumed following the addition of 5000 mg 1-t of calcium chloride. During batch studies utilising an acclimatised seed derived from settled sewage, Walker (1975) observed virtually no biodegradation of 1: I NTA complexes with cadmium and copper; chromium and nickel substantially reduced NTA biodegradation but iron, manganese and zinc had little or no effect. Similar studies carried out by Gudernatsch (1975) and Shannon et aL (t 978) confirm that the biodegradation of NTA is very dependent on the specific metal complex. Rossin et at. (1982c) studied an activated sludge pilot plant treating real settled sewage and demonstrated longer acclimatisation times at influent metal concentrations typical of mixed domestic/industrial sewage ("high") compared to influent metal concentrations typical of domestic only sewage ("low") (Table 5). Doubling the influent NTA concentration from 7.5 to 15mgl -~, at 9 days sludge age, reduced NTA removal from >99 to 93 and 5570 at "'low:' and "'high" influent metals respectively (Rossin e t a / . , 1983a). A similar effect was observed during transient temperature reductions (Stephenson et a / , 1983a). Incomplete biodegradation of NTA might be expected to have an adverse effect upon hea~y metal removal in sewage treatment processes since NTA is a strong chelating agent. Stoveland et aL (1979a) found reduced removals for copper, nickel and zinc in laboratory scale activated sludge units during the period before acclimatisation when the effluent NTA concentrations were greater than t.7 mgl . Similar studies on an activated sludge pilot plant also demonstrated high effluent concentrations of copper, lead, nickel and zinc associated with high effluent NTA concentrations (Rossin et al., 1982c). During the shock loading of NTA to laboratory scale activated sludge units already acclimatised to NTA, the effluent Table 4. Acclimatisation prior to NTA biodegradation in the concentrations of NTA were elevated to an average activated sludge process value of 4mg 1-~ (Stoveland et a/.. 1979b) and there Activated sludge Acctimatisationtime Reference system (day) Table 5. Acclimatisationtime for NTA m activated sludge at different influent metal concentratio!ls Swisher et al. ( 1 9 6 7 ) Laboratory 14-2 I Bouveng et al. (1968) Pilot >30 (Rossin et al., 1982c)

Janicke (1968) Pfeil and Lee (1968) Gudernatsch (1970) Shumate et at. (1970) Renn (1974) Stoveland et al. (1979a) Rossin et al. (1982c) Vashon et al. (1982)

Laboratory Laborator,, Laboratory Full scale Full scale Laborator', Pilot Laboratory

14-25 7 20 36 90 16-31 6-25 28

Sludge age [nfluent metal (day) concentration

Acclimatisation t~n~: (day)

4

Lov,

20

4 9 9

High Low High

> IS" 15 25

"No acclimatisation during the 18 days of ~,tud'~

NTA as a detergent builder was a significant reduction in the removal of lead, nickel and zinc. "'Wash day" simulations on an activated sludge pilot plant also resulted in elevated effluent concentrations of N T A and associated reductions in the removal of lead, nickel and zinc (Rossin e t al., 1983a). Nickel and zinc that had previously been adsorbed by the mixed liquor were mobilised to the effluent. At low sludge ages and " h i g h " influent metal concentrations, cadmium removal was also adversely affected. Nickel and zinc removals were reduced at effluent N T A concentrations as low as 0.7 and 0 . 6 m g l -~ respectively. During transient temperature reductions on laboratory scale activated sludge simulations decreases in the removals of cadmium, copper, lead and zinc were observed when the effluent N T A concentration was high (Obeng e t al., 1982). Cleasby e t al. (1974) studied the removal of cadmium, chromium, copper, iron, manganese, nickel and zinc in the presence of N T A in a pilot scale trickling filter. When acclimatised, only the removal of manganese was significantly reduced in comparison with the main plant. Studies on the effect of N T A on metal removal in full scale treatment plants have largely been confined to Canada. A short term detergent substitution study (Shannon, 1975) and a later study of 13 Canadian waste water treatment plants based on discrete sampiing (Woodiwiss e t al., 1979) were largely inconclusive, although the latter study did indicate some positive correlations between N T A and metals in the effluents. Wei e t al. (1979) described a more comprehensive study of a full scale activated sludge plant which indicated that of the four metals studied lower N T A removal generally resulted in reduced removals of aluminium, iron and zinc, and had some effect on copper. Effect of temperature

Bouveng e t al. (1968) reported removals of N T A in laboratory activated sludge units of only 2 5 ~ at 5°C. Eden e t al. (1972) achieved essentially complete biodegradation of influent N T A concentrations of 5 and 2 0 m g l -~ at 20°C but virtually no removal was observed at 5°C. Low temperatures also decreased the removal of N T A to a similar degree in aerated lagoons (Rudd and Hamilton, 1972; Rudd e t al., 1973) and trickling filters (Bouveng e t al., 1970; Cleasby e t a L , 1974). Obeng e t al. (1982) in experiments using laboratory scale activated sludge simulations demonstrated that there was no adverse affect on N T A biodegradation when the temperature was decreased from 17.5°C to approx. 12.5:C over 4 h, remaining at 12.5°C for a further 8 h. Two experiments with a reduction in temperature from 17.5 to 9.5°C over 8 h, remaining at 9.5~C for 6 h before returning to 17.5°C, resulted in decreases in N T A removal from 98 to 60~o and from 95 to 57~. A series of experiments using an activated sludge pilot plant treating settled sewage demonstrated that transient temperature reductions

261

Table 6. Effect of temperature on NTA removal during biological sewage treatment NTA Temperature removal Reference (C) (",,) Treatment Rudd and Hamilton (1972) 05 22 AL Bouveng et at. (1968) 5.0 25 ASL Eden et al. (1972) 5.0 0 ASL Rudd and Hamilton 11972) 5.0 47 AL Stephenson et at. (1983a) 6.0 79-85 AS Stephenson et al. (1983a) 7.5 90-98 AS Wei et al. (1979) 9.4 88 AS Obeng et at. (1982) 9.5 57-60 ASL Wei et at. (1979) 9.8 85 AS Eden et at. (1972) lif0 94 ASL Cleasby et at. (1974) 10.0 82 TF Stephenson et al. (1983a) lif0 93-100 AS Wei et al. (1979) 10.3 58 AS Obeng et al. (1982) 12.5 85 ASL Stephenson et al. (t983a) 12.5 >95 AS Wei et at. (1979) 14.5 89 AS Rudd and Hamilton (1972) 15.0 93 AL Wei et al. (1979) 15.5 99 AS Stoveland et al. (1979a) 17.5 83-100 A S L Obeng et al. (1982) 17.5 95-98 ASL Eden et al. (1972) 20.0 >98 ASL Bouveng et at. (1968) 25.0 80-85 ASL AL--aerated lagoon; ASL--activated sludge (laboratory): AS-activated sludge: TF--trickling filter. from 17.5 to 10~C caused a slight decrease in N T A removal (Stephenson e t al., 1983a). The decrease in N T A removal was more significant during temperature reductions from 17.5 to 7.5~C and reductions from 17.5 to 6°C resulted in a substantial decrease in N T A removal. Analysis of data for a full scale activated sludge plant revealed a significant difference in N T A removal between 15.5 and 9.4~C and between 14.5 and 10.3~C for inftuent concentrations of 2.5 and 20 mg 1- ~ respectively (Wei e t al., 1979). The effect of temperature on N T A removal during biological sewage treatment is summarised in Table 6. Effect

of NTA

loading

Results from an activated sludge pilot plant study demonstrated that the influent N T A concentration affects the acclimatisation time (Rossin e t al., 1982c). The higher the influent concentration, the longer the acclimatisation time required (Table 7). Furthermore, when activated sludge is fully acclimatised, the percentage removal of N T A has been shown to be influenced by its influent concentration. With an influent N T A concentration of 2 0 m g 1-~. Janicke (1968) observed 98~o removal of N T A , while only 70~o removal was achieved at 50 mgl-~. Shumate e t al. (1970) obtained 9 0 ~ removal of N T A at influent Table 7. Acclimatisation time for NTA in activated sludge at different influent NTA concentrations (Rossin et aL. 1982c) Sludge NTA influent Acclimatisation age conc. time (day) (mg l-I) (day) 4 7.5 15 4 15.0 20 9 7.5 12 9 15.0 16

2~2

R. PERRY et al.

Table 8 Removalof NTA in activated sludge at different influent NTA concentrations(Wei et el. 1979) 3,~erage influent Average Average NTA NT-X.concentration temperature removal fmgl-'l l C) io,) )7 .... ;" ........ 88.3-- 6.t 9= 75.9 145 9._" 64.8 21.9 I0.3 58.3

concentrations of 5.9 mg l-~ and only 75~.; removal at 11.7 mg l -~. Wei et al. (1979) found that the overall removal rate of NTA in a full scale wastewater treatment plant decreased with increasing inftuent NTA concentration (Table 8). This effect was statis.tically significant when comparing removal at influent NTA concentrations of 2.7 and 14.5mgl -~ when temperatures were similar (Table 8). Variations in influent NTA concentration, as would be expected in full scale works, have also been demonstrated to influence NTA removal. Shumate et al. (1970) increased the influent NTA concentration to an activated sludge plant from 1.5 to 5.9mgl -~ and the removal efficiency decreased from about 90 to 60".0, a period of one week elapsing belbre 90')/o removal was resumed. Cleasby et al. (1974) reported no reduction in NTA removal by a trickling filter when the influent NTA concentration was increased from 2.6 to 6.1 mg 1-~ although several days acclimatisation were required when the NTA concentration was further increased to 11.6 mg l -~. Stoveland et al. (1979b) doubled the influent concentration of N T A from 10 to 20mgl -~ in a series of three experiments and the effluent NTA concentration rose from near zero to between 3.6 and 4.4 mg 1- ~. Obeng et aL (1981) using two laboratory scale activated sludge units doubled the influent NTA concentration from 10 to 2 0 m g l -~ for a period of 12h and the effluent concentration of NTA increased from 3 to 10 mg I-~ and from less than detectable to 5 mg lrespectively. Three sludge ages (4, 9 and 12 days), two influent metals concentrations ("low" and "high") and two influent NTA concentrations (7.5 and 15 mg 1 ~) were selected to study the influence of simulated "'wash-day" shock loading on the biodegradation of NTA in an activated sludge pilot plant (Rossin et al., 1983a). The effect appeared to be independent of initial influent NTA concentration. In general, experiments at low sludge ages and "high" metal concentrations had the greatest effect on NTA removal. Effect o f water hartbwss

Stoveland et al. (t979a) observed no biodegradation of influent NTA after 60 days in soft water laboratory scale activated sludge units. In units supplied with hard water synthetic sewage the acclimatisation period varied between 16 and 31 days. Vashon et al. (1982) achieved near complete biodegradation of NTA in both hard and soft water laboratory scale activated sludge units after approx.

28 days. The rate at which full acclimatisation was achieved was slower in the soft water units compared tO the hard water units. The different observations of Stoveland et al. (1982) and Vashon et al. (t982) in achieving acclimatisation in soft water are probably attributable to differences in operating parameters. The units of Stoveland et al. (1979a) treated an influent containing 10 mgl-~ NTA and the metals cadmium, chromium, copper, lead, nickel and zinc whilst Vashon et al. (1982) used an influent containing 8 mg 1-t NTA and no additional metals except for iron. A higher concentration of influent NTA will increase the acclimatisation time of an activated sludge plant (Rossin et al., 1982c). Moreover, the iron-NTA complex is readily degradable whereas NTA complexes with cadmium, copper, nickel and zinc are less easily degradable (Warren, 1974). Bj6rndat et al. (1972) observed that the biodegradation of the copper-NTA complex decreased from 96 to 78Vo by decreasing the water hardness from 276 to 29 mg 1- Kas calcium carbonate. Walker (1975) also observed that decreasing the water hard, ness from 240 to 120 to 0 mg 1- ~as calcium carbonate reduced the overall biodegradation of a 2:1 copper-NTA mixture from 93 to 51 to ~I" rEspectively. A subsequent addition of iron increased biodegradation in both hard and soft waters. Effect o f p r o c e s s p a r a m e t e r s

Rossin et al. (1982c) investigated the effect of the mean cell residence time (or sludge age) of an activated sludge pilot plant upon the time required for acclimatisation to NTA. Acclimatisation was more rapid at higher sludge ages when the mixed liquor suspended solids were higher and the wastage rate lower (Table 9). The sludge age and influent heavy metal concentrations were interactive as acclimatisation did not occur at low sludge age (4 days) and "high" influent metal concentrations. Once acctimatisation has occurred, sludge age and mixed liquor suspended solids concentration have an effect on NTA removal. Changes in the inftuent NTA concentration from l0 to 20 mg 1-~ for 12 h were studied in acclimatised laboratory scale activated sludge simulations at two different mixed liquor suspended solids concentrations (Obeng et al., 1981). In the simulation with the lower mixed liquor suspended solids concentration (2054 mg 1- t), the effluent NTA concentration increased from 3 to 10 mg l -~. In the simulation with the higher mixed liquor suspended solids concentration (4330 mg I -~) the effluent NTA concentration

Table 9. Acclimatisationtime for NTA in activated sludge at different sludge ages (Rossin et al.. 1982e) Sludge age MLSS Acclimatisation (day) (mgl-I) (day) 4 1514 15 9 3049 I2 12 3536 6

NTA as a detergent builder increased from tess than detectable to 5 mg 1-~. An increase in influent NTA concentration from 15 to 30 mg I -~ was applied to an acclimatised activated sludge pilot plant with a "'high" influent metals concentration (Rossin et al., 1983a). Effluent NTA concentrations increased to maximum values of 1.7, 11.1 and 0.7mgl -~ at 4, 9 and 12 day sludge ages respectively. The same series of experiments was undertaken at "'low'" influent metals concentrations and an increase in influent NTA concentrations from 7.5 to 15.0mg 1-]. Effluent NTA concentrations increased to maximum values of 2.0, 0.60 and 0.60mgl -~ at 4. 9 and 12 day sludge ages respectively. Gudernatsch (I 970) found that decreasing the hydraulic retention time of laboratory activated sludge units receiving 20-30 mg 1- t of NTA from 6 to 3 h resulted in a reduction in NTA biodegradation from near complete removal to approx. 80% removal for 4 days prior to recovery. At an influent NTA concentration of 200 mgl-~ a similar reduction in retention time had little effect on NTA removal efficiency (Swisher et al., 1967).

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Results o f f i e l d studies

Several studies of NTA removal in biological wastewater treatment have been undertaken on a "field scale". These are summarised in Table 10. The removal of NTA in trickling filters has been studied on a pilot scale (Cleasby et al., 1974) and full scale plant (Bouveng et al., 1970). Both used composite sampling regimes which do not reveal short term variations in NTA removal. Following an initial acclimatisation period of 2-3 weeks Cleasby et al. (1974) observed approx. 90~o removal of NTA by the trickling filter serving a population equivalent of only 60. Further acclimatisation periods were necessary after the wastewater flow had ceased for 4 days and on another occasion when the addition of NTA was interrupted for 25 days. Reduced NTA removals were observed in cold weather and sporadically for no apparent reason. Bouveng et al. (1970) reported that the removal of NTA in a plant serving a population equivalent of 12,500 exceeded 80~o in most instances during summer and early autumn. Weekly figures for removal indicated poor NTA degradation at increased hydraulic loads and low temperatures. Rudd et al. (1973) studied the removal of NTA by an aerated lagoon and an activated sludge plant in Canada. The size of each treatment system or the population equivalent served were not stated. Monthly discrete samples of effluent were analysed for NTA but influent NTA concentrations were not quoted due to deficiencies in the analytical technique. An activated sludge plant removed influent NTA concentrations of 1.5, 5.9 and 11.7 mg 1-~ with average efficiencies of 89.4, 90.0 and 75.20./o respectively (Shumate et al., 1970). The plant was underloaded as the average flow received was 0.4 mg day- ~compared to a design flow of 1.0 mg day -~. Composite samples

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were collected over 24 h despite no NTA being dosed to the plant over 10h of the night. Some daily removals of NTA were less than 50°0 during cold periods. Renn (1974) studied an activated sludge plant serving a population of only 465 with a high average aeration retention time o f 20 h and received virtually no flow after midnight. Even after acclimatisation some daily effluent NTA concentrations exceeded 7.3 mg 1-'. A 3 week detergent substitution study on a plant serving 400 people was carried out by Shannon (1975). Sampling during the last week revealed an average influent NTA concentration of 2.19mgl -~ with an overall removal of 60.6°'o at an average temperature of 14.1:C. There was no evidence of adverse effects on metal removal, but as the authors stated "the substitutions were of too short a duration to provide definite conclusions". This comment is probably equally applicable to NTA removal. An activated sludge plant with a 20 h retention time treating wastewater from a population equivalent of 5500 removed NTA an average 9 0 - 9 5 ~ during the summer and only 50~o in winter (Gudernatsch, 1974). Wei et al. (1979) studied the removal of NTA by an activated sludge plant receiving wastewater from a population equivalent of 2000. The plant required approx. 2-3 weeks to acclimatise to higher influent NTA concentrations. When acclimatised, removal of NTA decreased with higher influent NTA concentrations and lower wastewater temperature. Woodiwiss et al. (1979) analysed random monthly discrete samples from 13 wastewater treatment plants in Canada. Therefore any changes in NTA concentration due to "wash days" or other factors have not been identified although inftuent concentrations of NTA recorded at activated sludge plants ranged from 0.06 to 15.75 mgl -~ and at trickling filters ranged from 0.02 to 19.69 mg I-L Concentrations of NTA in the effluent from activated sludge plants were as high as 6.40 mg 1- ~and from trickling filters the peak value recorded was 7.68 mg l -I. Hardly any of these so called "'field scale" studies reflect the probable situation in Europe if NTA were to be introduced. In most cases the influent NTA concentrations were lower than would be expected in Europe (Rossin et al., 1982a). As a consequence of the composite sampling regimes emPloyed (Table 10) any large variations in influent and effluent NTA concentrations within a day were not revealed. At any time during the composite sampling period the concentrations of NTA in the effluent may have reached a value high enough to cause mobilisation of heavy metals to the effluent (Rossin et al., 1982c, 1983a). It is unfortunate that in the only recent field study in Europe (Oude, 1982) NTA was only incorporated into 6~o of the wash formulations in use and as a result median influent NTA concentrations to the sewage treatment works were between 0.080 and 0.254 mg 1-~. Moreover the use of a composite sampling protocol in this study again prevented any observations of transient variations in NTA removal.

SEPTIC TANK TREATMENT

It has been estimated that 5'!11 of the U.K. population is served by septic tanks and cesspits, although the distribution between these processes has not been specified (Department of the Environment and National Water Council, 1981). In the United States 5.6°,0 of the population or 12.944 communities were served by septic systems in 1980 and 38.6
BEHAVIOUR OF NITRILOTRIACETIC ACID IN RECEIVING WATERS

The pollution of receiving waters and in particular rivers has become the subject of considerable concern (Water Research Centre, 1976). In industriatised countries where water resources are limited, sewage effluents may contribute a substantial part of the volumetric flow of lowland rivers. At certain times of the year in some localities in the U.K. sewage effluent may represent 90~ of the flow in surface waters, some of which are later abstracted for drinking water (Hawkins et al., 1980) and approx. 30°0 of the U.K.'s water supplies contain some sewage effluent (Fielding and Packham, 1977). tn certain areas of Europe raw sewage is discharged to inland water bodies (Haddrilt et al., 1983) while raw and treated sewage are discharged to estuarine and coastal waters.

NTA as a detergent builder Inland waters

The biodegradation of NTA in receiving waters has largely been studied in batch laboratory experiments although Canadian monitoring data has been reported ( R u d d e t al., 1973; Shannon et al., 1974: Woodiwiss et al., 1979). Bacteria capable of utilising NTA have been isolated from lake and river waters in Sweden (Forsberg and Lindquist, 1967). Warren and Malec (1972) demonstrated the aerobic biodegradation of NTA in previously acclimatised and unacclimatised river waters, but were unable to confirm complete removal due to insensitivity of the analytical method employed. Batch tests by Thompson and Duthie (1968) also demonstrated biodegradation of NTA in river water after an acclimatisation period. A reduction in the rate of NTA biodegradation with decreasing temperature in batch studies has been reported by several authors (Shannon et al., 1974; Bott et al., 1980; Larson et al., 1981). Swisher et al. (1973) found that the extent of NTA removal in river water decreased with increasing metal concentration in the case of cadmium, copper, lead, mercury, nickel and zinc while the concentration of iron(Ill) had no effect on NTA biodegradation. Complexes of NTA with copper, mercury and nickel in lake water did not biodegrade over a time period exceeding 100 days while cadmium and lead-NTA complexes biodegraded in 60 and 25 days respectively (Chau and Shiomi, 1972). Complexes of atuminium, calcium, chromium, iron, magnesium, manganese and zinc biodegraded in 17 days or less. It would appear that the effects of temperature and heavy metal concentration upon the biodegradation of NTA in river water are similar to those observed during biological waste water treatment. Larson et al. (1981) also demonstrated a decreased rate of NTA biodegradation in river water at low dissolved oxygen concentration. Although Larson and Davidson (1982) observed acclimatisation and biodegradation of NTA at low initial concentration (0.005 mg l -t) in batch studies on river water it is probable that natural waters would always contain trace quantities of NTA, even under optimal conditions for biodegradation, due to the continuous discharge of NTA into receiving streams (Tiedje, 1980). Rudd et al. (1973) could not detect NTA in rivers receiving effluent containing NTA in summer or winter, although the extent of dilution of the treatment works effluent by the receiving stream was not stated and the analytical method used only had a detection limit of 0.02 mg 1-~. Shannon et al. (1974) reported a field study during which NTA concentrations were monitored over a year downstream of a sewage effluent discharge. Under winter conditions (0.5-3°C) the mean NTA concentration was 0.106rag I -~ approx. 0.8 km downstream from the input which compares with summer ( > 10°C) concentrations of 0.01 mg 1 -t or less. A survey in Canada, when NTA constituted on average 6~o of detergent compositions, revealed

265

that geometric mean concentrations of NTA in receiving streams above and below effluent discharges ranged from 0.002 to 0.021 mg 1- ~and from 0.002 to 0.045 mg 1-~ of NTA respectively (Woodiwiss et al.. 1979). When the average NTA content of detergents was 15°4 the geometric mean concentrations of NTA in receiving streams above and below effluent discharges ranged from <0.001 to 0.028mgl -~ and from < 0.001 to 0.667 mg l-~ respectively. The results from Canada confirm that the introduction of NTA detergents would result in the persistence of variable concentrations of NTA in receiving streams. Estuarine a n d coastal waters

The persistence of NTA in estuarine and coastal waters will be influenced by a combination of complex chemical, physical and biological factors (Bartholomew and Pfaender, 1983). Using radiolabelled NTA, Bourquin and Przybyszewski (1977) were unable to isolate estuarine bacteria capable of metabolising NTA in estuarine environments but were able to isolate estuarine bacteria capable of metabolising NTA in freshwater. Their data indicated an interference with NTA catabolism by some unknown factor of the estuarine environment rather than an absence of potential NTA-degrading bacteria. It has been suggested that the high ionic strength and the chloride ions in such waters may inhibit monooxygenase activity necessary for NTA biodegradation (Environmental Protection Agency, 1980). Kirk et al. (1983a) demonstrated that a bacterial population of marine origin was unable to degrade NTA under aerobic or anoxic conditions in a laboratory simulation of sewage disposal to sea. A study by Bartholomew and Pfaender (1983) demonstrated the maximum rate of NTA utilisation (vm,,) to be 500 ng I - ' h -~ at 12~C and 56 ng 1-t h -t at 7°C in freshwater, 2 0 0 n g l - ~ h -~ at 14:C and 175ngl-~h-~ at 6°C in estuarine water and 8 5 n g l -t h -~ at 8~C in seawater. However. the authors acknowledge that these rates are strongly influenced by spatial and temporal variations in the environment and do not correlate consistently with general characteristics of the microbial community, direct measurement of biodegradation rates being preferable. The presence of NTA in estuarine and coastal waters may result in significant effects on the metabolism and growth of planktonic algae in the marine environment. The available literature suggests that NTA might effect algal growth by three major mechanisms: (i) supplementing, upon degradation, the nitrogen requirements of phytoplankton particularly in nitrogen limited waters; (ii) increasing the availability of limiting trace elements via chelation; (iii) reducing the availability of toxic metals through chelation and thereby stimulating the growth of those species previously inhibited.

26¢,

R. PERRY et al.

The inability of bacteria and phytoplankton to significantly metabolise NTA in saline waters coupled ~vith the relatively low contribution of NTA to selvage nitrogen, suggests that nitrogen enrichment by NTA is unlikely. Graneli and Edler (1983) and Kirk et al. (1983a) conclude that the most significant effect of NTA in coastal waters would be related to its metal chelating ability. Yentsch et al. (1974) reported increased growth of a marine dinotlagellate Gonya,dax tamarensis, which is associated with toxic "'red tide" blooms, in the presence of 10 mg 1-~ NTA. Data presented by Anderson and Morel (1978) suggest that the growth of G. tamarensis may be totally inhibited by copper toxicity in natural seawater under conditions that leave other algae relatively unaffected and that organic chelation of this copper may be necessary before the cells can successfully multiply to bloom proportions and were able to demonstrate this affect using EDTA. A more recent study by Gran~li and Edler (1983) demonstrated the stimulatory effect of 0.0188mgl -~ NTA on the dinoflagellate Prorocentrum minimum and the diatom Skeletonema costamm, and ascribed this to a reduction in copper (ll) toxicity. The effect of NTA was found to exceed that of EDTA.

BEHAVIOUR OF NITRILOTRIACETIC ACID DURING SLUDGE TREATMENT AND DISPOSAL

Of a total sludge production of 35 million tonnes per annum (wet wt) in the United Kingdom, approx. 50% is digested (Collinge and BruCce, 1981), heated anaerobic digestion being the most common stabilisation process (Department of the Environment and National Water Council, 1981). Despite the importance of anaerobic systems for sludge treatment (as well as for the treatment of sewage in septic tanks) few studies have been undertaken to determine the behaviour of NTA under such conditions. Enfors and Molin (1971) reported the ability of an isolated bacterial strain to utilise NTA as a sole carbon source with nitrate as electron acceptor. The extent of degradation was found to be variable (Claesson, 1971). Enfors and Molin (1973a, b) detailed the isolation of a facultative anaerobe capable of degrading NTA anaerobically, mud samples from the Stockholm area being utilised as inoculum for anaerobic enrichment cultures in continuous and batch systems. The lag phase prior to any measurable growth was found to vary between 8 and 27 days, much longer than the lag for aerobic cultures. Furthermore, no growth was observed in media containing NTA as the sole nitrogen source. Effect of anaerobic sludge digestion Thompson and Duthie (1968) operated three laboratory anaerobic digesters on a synthetic waste and reported no effect on gas production and pH for an NTA feed of between 3.7 and 66mg 1-~. No other

operational parameters were reported. Klein (1974) described a s~,stem involving laboratory digesters with influent NTA concentrations of I1, 22 and 44 mg 1-t operating on raw sludge with a retention time of 30 days. Nitrilotriacetic acid recovery in the effluent averaged 92, 86.4 and 92~0 respectively demonstrating that no biodegradation occurred. Kirk et al. (1982a) observed removals of between 29 and 45,°.0 of NTA in digesters receiving co-settled primary and waste activated sludge, unacclimatised to NTA, over a period of 120 days (6 retention times). The addition of sodium azide and subsequent failure of digestion demonstrated that this removal was not the result of biological activity while batch tests indicated adsorption onto the sludge solids as being a significant removal mechanism. Moore and Barth (1976) described a laboratory digester which was operated for a period exceeding one year. The first month acted as a control experiment with no NTA additions. With a feed of primary sludge containing 14.6 mg 1-~ of NTA no appreciable removal was observed. Similar results were obtained for a feed containing 80% (v/v) primary sludge and 20% (v/v) waste activated sludge previously acclimatised to NTA. However, when the proportion of waste activated sludge was increased to 50% (v/v) the NTA concentration was found to gradually decrease to zero. This observation was explained by the presence of facultative bacteria in the activated sludge capable of degrading NTA anaerobically. Stephenson et al. (1983c) confirmed these observations by operating anaerobic digesters on mixed primary sludge containing 15 and 30% waste activated sludge acclimatised to NTA. Following a lag phase of between 4 and 16 days influent NTA concentrations of 10 and 30 mg 1-~ were removed by a mechanism which was concluded to be biological. Co-settling of surplus activated sludge in primary tanks is not universal practice. Land disposal o f sludge At present the most common options for sewage sludge disposal are dumping at sea, landfill of dewatered sludges and the application of sludge to agricultural land (Sterritt and Lester. 1980). In 1977, 67% of the sludge produced in the United Kingdom was disposed of to land, 29~o to sea and 4% was incinerated (Department of the Environment and National Water Council, 1981). The fate and effects of NTA in sludge or effluent applied to land has been the subject of only very limited study. Tiedje and Mason (1971) reported the ability of soil micro-flora to degrade NTA aerobically, removals being greater in sludge-amended soils, and confirmed this observation by incubating sterile and viable soil samples aerobically with NTA (Tiedje and Mason, 1974). Reported removals in the latter study ranged from <5% in some sub-soil samples up to 80% with no correlation evident between removal and pH, drainage, texture or plant cover. Sludge amended

NTA as a detergent builder soils exhibited complete NTA dissimilation after a tag phase of several days whilst mineral soils failed to exhibit a lag phase but displayed more variable degradation. Perfusion systems indicated that NTA degradation would only occur in the presence of oxygen (Tiedje and Mason. 1974) and pure culture enrichments isolated a Pseudornonas species from soil which was capable of aerobically utilising NTA as a sole nitrogen or sole carbon and nitrogen source (Tiedje et al., 1973). The production of N-nitrosoiminodiacetate was demonstrated in a recycling soil percolator by micro-organisms growing in mixed culture with NTA as sole carbon source and sodium nitrate as nitrogen source (Pickaver. 1976) while this did not occur in pure culture (Pickaver, 1974). Tabatabai and Bremner (1975) reported that the aerobic and anaerobic decomposition of NTA in soil incubations under viable conditions was terminated by autoclaving. They were unable to explain the rapid anaerobic decomposition, which was contradictory to the findings of Tiedje and Mason (1971) who failed to demonstrate degradation under anaerobic conditions. Means et al. (1980) studied the degradation of a 2 x 10 -3 M solution of NTA with soil under air and nitrogen. Degradation rate was found to increase in the presence of nutrients and decrease with decreasing oxidation-reduction potential, although no truly anaerobic conditions were investigated. The authors caution against the extrapolation of results from such idealised laboratory experiments to the natural environment (Means et al., 1980). Dunlap et al. (1971, 1972) investigated the fate and effect of NTA in soil profiles by applying a weak synthetic sewage containing 36.6mgl -~ NTA to sandy, loam and clay-loam soil columns under saturated and unsaturated conditions. Removal was near complete in unsaturated columns, but was poor in saturated columns, suggesting that aerobic biodegradation could occur, but not anaerobic biodegradation. Metal enriched sand columns irrigated with a similar solution exhibited metal mobilisation although the lack of adequate controls made interpretation of the results difficult. Sorption was greater by loam soil than sand and was deemed sufficient to retard but not prevent the percolation of NTA to groundwater. Studies with model aquifers indicated that NTA would degrade slowly in the essentially anaerobic groundwater environment (Dunlap et al., 1971, 1972). Klein (1971, 1974) operated soil columns and pilot scale percolation fields under unsaturated and saturated conditions and obtained similar results, removal being nearly 95~ in aerobic soil columns and 10-15% for anaerobic columns at concentrations up to 73.2 mg 1-~ of NTA. Kirk et al. (1983b) reported that surface applications of sewage sludge to land, consistent with the spreading of sludge on grazing land, would result in significant quantities of NTA passing through the topsoil and possibly entering ground waters, while sludge/soil mixtures typical of sludge application to arable land followed by plough-

2~,-

ing, would retain a greater proportion of the applied NTA in the topsoil. The most significant NTA removal mechanism over the short term was concluded to be adsorption, which was consistent with the results of batch adsorption studies. Hrubec and Van Delft (1981) investigated the behaviour of NTA during artificial groundwater recharge of river water in two pilot plant scale sand filters. At a hydraulic retention time of 6 days NTA appeared in the effluents after 2 weeks, increased to 56 and 81°o of the influent concentration of 0.73 mg 1-~ over the following 2 weeks and then gradually decreased as the result of aerobic biodegradation. There have been few reports on the effects of NTA on metal uptake by plants. Abdulla and Smith (1963) observed slight increases in the iron, calcium, magnesium and zinc contents of cabbage plants grown in soils treated with 1000mgl -~ of NTA, the effect being more marked in alkaline soils. Wallace et al. (1974) reported increased concentrations of copper, zinc, manganese and nickel in the leaves of soybean plants grown in a calcareous loam soil containing 100-1000 mg 1-~ of NTA, although bush bean plants grown in an acidic soil (pH 6) exhibited more variable uptake. Sludge d u m p e d at sea

Sludge disposal to sea in the U.K. accounted lbr 29~o of sludge production in 1977 (Department of the Environment and National Water Council. 1981), disposal being concentrated at a few specific sites, particularly the outer Thames Estuary which receives 45yo of the total dumped at sea (Fish, 1983). Concern has been expressed over the recalcitrance of NTA in saline systems (International Joint Commission, 1978) especially with regard to the ocean dumping of sewage sludges contaminated with NTA (Environmental Protection Agency, 1980). The fate of NTA in the marine environment has only been superficially examined. An assessment of the toxicology of NTA concluded that it may present a significant hazard to marine teleosts and higher invertebrates (Eisler et al., 1972). Erickson et al. (1970) reported the utilisation of NTA by marine bacteria but determined growth by visual turbidity following an addition of 500 mg l-t of NTA as sole nitrogen or nitrogen and carbon source. Neither bacterial viability nor NTA utilisation were determined. Bacteria have been isolated from seawater capable of metabolising NTA as a sole nitrogen or nitrogen and carbon source in synthetic estuarine medium, but NTA utilisation in this system was poor (Erickson et al., 1970). Thin layers of deposited sludge appear to undergo aerobic decomposition while anaerobic conditions can occur in the lower layers of sediment which are more than 10mm thick (Grunseich and Duedall, 1978) and anoxic sediments have been observed surrounding a sewage outfall to sea (Morel et al., 1975). Kirk et al. (1983a) demonstrated that a bacterial population of marine origin was unable to degrade NTA under

2~,:~

R. PEp,~y et al.

either aerobic or anoxic conditions. NTA concentrations of 0.5 and 7.5 mg t-~ were investigated, NTA representing the major carbon and nitrogen source at the higher concentration. It was concluded that NTA would persist or accumulate in the immediate vicinity of sludge dumping sites and points of sewage discharge to coastal waters (Kirk et al,, 1983a) with possible consequences for metal speciation and bioavailability (Niirnberg and Raspor, 1981). It is generally accepted that the physico-chemical state of trace metals in seawater significantly affects their potential to enter marine food chains, their transport and their accumulation in sediments (Musani et al., 1980). It has been shown that the discharge of NTA to the marine environment could have adverse effects on heavy metal accumulation in marine food chains (International Joint Commission, 1978) and may alter the pathways of heavy metals in the estuarine environment (Jones, 1978) in addition to the promotion of "red tide" blooms in coastal waters via a reduction in toxic copper(ll) concentration (Graneli and Edler, 1983). ENVIRONMENTAL MONITORING Ot= N1TRILOTRIACETIC ACID

Monitoring of NTA is important both environmentally and during its manufacture and distribution. Despite the wealth of literature documenting the aerobic biodegradation of NTA natural waters would always contain detectable concentrations even under optimum conditions for biodegradation (Tiedje, 1980; Larson and Davidson, 1982). A prerequisite for the determination of NTA in environmental samples is an analytical procedure applicable to the particular matrix involved, which is both accurate and reliable in the concentration range of interest. Analysis

Methods for the determination of NTA in waters and wastewaters have been reviewed by Kirk and Lester (1981) who concluded that the most frequently applied analytical techniques were gas-liquid chromatography, polarography and colorimetry. An interlaboratory comparison of analytical techniques was undertaken by the Procter & Gamble Co. and the Canada Centre for Inland Waters using gas-liquid chromatography (Aue et al., 1972) and polarography (Afghan and Goulden, 1971) respectively, as part of a national Canadian monitoring programme. The results of the comparison were summarised by Matheson (1977) who stated that the "variability found in the comparison samples clouds the interpretation of the monitoring data which are based on spot samples taken at considerable time intervals". At concentrations exceeding 0.010mgl-~ the Procter & Gamble values were significantly higher, the differences possibly being due to contamination, differences in the species being identified and changes during storage (International Joint Commission, 1978).

Severe positive interference has been reported during the analysis of NTA in settled sewage and sewage effluent by gas-liquid chromatography, and significant interference in potable water at low NTA concentrations (Kirk et al., 1982b). Aue et aL (1972) and Games et al. (1981) observed similar interference, the latter applying radiolabelled ['aC]NTA to overcome variable NTA recovery from sewage influent and effluent samples. Interference encountered by Kirk et al. (1982b) was not reduced by the anion exchange elution modifications outlined by Games et al. (1981), nor by the separate application o1" two internal standards, dibutylphthalate (Bott et al., 1980) and nitrilotripropionic acid (Games et al., 1981). A nitrogen specific detector has been applied to increase resolution (Williams et al., 1977) although Games et al. (1981) reported wide fluctuations in response from this type of detector. Differential pulse polarography has been concluded to be the most precise, accurate and rapid method for the determination of NTA in a range of waters and wastewaters (Kirk et al., 1982b). Recovery for this technique may be improved by preparing standards in a matrix of similar composition to the sample, but uncontaminated by NTA (Kirk et al., 1982b). M o n i t o r i n g studies

Several major monitoring programmes have been undertaken in Canada and the United States (prior to the voluntary suspension of the use of NTA in U.S. detergents) in order to investigate the occurrence of residual NTA concentrations or detect environmental accumulation, but none included a thorough study of NTA recovery and the effect of interferents on the concentrations determined. Thayer and Kensler (1973) reported concentrations of between 0.025mgl -~ (detection limit) and 0.125mgl ~ of NTA in 6 out of 279 samples of potable water in the United States although this study was of limited value since only 6-10 °/ of households in the area monitored were using or had been using a detergent containing NTA, Canadian monitoring data reported by Matheson (1977) were based on discrete samples taken at monthly or 3 monthly intervals and would not be representative of diurnal or seasonal variations in NTA concentrations. Potable water concentrations were generally below 0.010 mgl-~ of NTA, although the number of samples exceeding this concentration appeared to be increasing annually throughout the monitoring programme. A study of 13 Ontario cities carried out as part of this programme was detailed by Woodiwiss et al. (1979). Monthly discrete samples of sewage influent, effluent and receiving streams above and below outfalts were analysed between January 1972 and March 1975. Concentrations were reported as geometric means together with the range. During the period April 1973 to March 1975, when NTA constituted 15°i, of detergents in Canada, 40go of the receiving streams monitored contained >0.037 mgl-~ NTA while 9",,

NTA as a detergent builder contained >0,183 mgl -t NTA. Estimated dilution ratios (stream effluent) ranged from 1.3 to 28.000 and the treatment works monitored served populations of up to 63,000. Malaiyandi et al. (1979) reported the results of a national survey of potable water undertaken by the Canadian Environmental Health Directorate. Concentrations of N T A in samples of treated water collected from 70 municipalities ranged from the detection limit (0.2,ugl -~) or below up to 0.1013 mg I -~. Recovery at or below 0,025 mg 1-~ was reported to be 94.7 + 4.20,/0 (8 samples), using the gas-liquid chromatographic method of Aue et al. (1972) as modified by Williams et al. (1977), DISCUSSION

Aerobic biological processes are essential for the effective removal of NTA during sewage treatment. Adsorption of NTA by biological solids as a removal mechanism is unimportant, with the possible exception of circumstances when NTA concentrations are too low for enzyme induction, since biodegradation of NTA will commence after a period of acclimatisation. In most studies this acc[imatisation period has been determined in laboratory and pilot scale activated sludge simulations under optimum conditions and has ranged from 6 to 90 days (Table 4). It is significant that the longest reported acclimatisation periods of 36 and 90 days were the only studies on full scale treatment works (Shumate et al., 1970; Renn, 1974). In each case this acclimatisation period occurred when the influent NTA concentration was increased from a base level of approx. 0.5mgl -~. The retention time (18h) of the plant studied by Shumate et al. (1970) was higher than that for conventional activated sludge plants serving the majority of the population in Europe and the influent concentration of 2 mg 1-~ NTA was much lower than the 10-15 mgl-~ anticipated in Europe which would be higher in hard water areas (Rossin et al., 1982a). As a consequence of the high per capita water usage, many of the field studies undertaken in Canada have reported low influent NTA concentrations (Table I0). A study of the 13 municipal treatment works in Canada when the average NTA content of detergents was 15~ exhibited a median influent concentration of 2.34 mg 1-~ (Woodiwiss et al., 1979). It is unfortunate that in a recent monitoring study in the Netherlands, NTA was only in very limited use in detergents (6~), and therefore exhibited a maximum value in the influent raw sewage of only 0.712mgl -t (Oude, 1982). It is possible that at such low influent NTA concentrations removal could have been solely due to adsorption (Stephenson et al., 1983b). This would not reflect the expected situation were NTA to be introduced on a large scale in Europe when its removal would be dependent upon the more sensitive biodegradation mechanism. Under ideal conditions in the activated sludge process with a steady influent NTA concentration the

26~1

length of the acclimatisation period is proportional to the NTA concentration, the greater the NTA concentration the longer the time required to achieve near complete biodegradation (Rossin et al., [982c). Indeed, the longest reported acclimatisation time of 90 days under field conditions was for an increase in influent NTA from near zero to 21.9 mg 1-t (Renn, 1974). The acclimatisation period and subsequent biological removal to NTA may also be affected by water hardness (Walker, 1975; Stoveland et al., 1979a). It has been proposed that biodegradation of m e t a l - N T A complexes proceeds via the most readily biodegradable complexes e.g. calcium-NTA (Warren, 1974). Once the calcium-NTA complex has been biodegraded the calcium ions released will substitute the metal ions present in less readily biodegradable m e t a l - N T A complexes, e.g. cadmium-NTA, so enabling the NTA to be biodegraded. It is probably for this reason that the presence of a high level of water hardness, and therefore calcium ions, can facilitate the biodegradation of NTA compared to the same system in soft water (Stoveland et al., 1979a: Vashon et al., 1982). All studies of the effect of water hardness upon the biodegradation of NTA during aerobic biological treatment have been made at laboratory scale, despite the statement by Vashon et al. (1982) that " . . . NTA is removed in real sewage treatment plants under a wide range of conditions including water hardness (Woodiwiss et al., 1979)". Woodiwiss et al. (1979) did not mention water hardness nor levels of calcium and magnesium in the sewage influent and effluent and receiving streams. Transient high influent concentrations of NTA, such as those occurring during "wash days", will result in a decrease in NTA removal during biological sewage treatment (Obeng et al.. 1981; Rossin et al., 1982b). Rossin et al. (1982a) observed a peak in the influent concentration of condensed phosphates, largely attributable to detergent builders, of up to four times the average concentration. A field study by Wei et al. (1979) confirms the existence of such wash day peaks for NTA when used as a detergent builder. Such studies have also confirmed reduced NTA removals at higher influent NTA concentrations (Shumate et al., 1970; Wei et al., 1979). Biodegradation of NTA is also adversely affected at low temperature. Laboratory and field studies indicate that below 10*C overall removal is reduced (Table 6) and below 7°C biodegradation is seriously impaired (Stephenson et al., 1983b; Van't Hof et al., 1983). During cold weather NTA removal in sewage treatment plants will be reduced and increased concentrations of NTA in the effluent will be discharged to receiving waters. Concentrations of NTA in effluent equal to or greater than 0 . 6 m g l -t could cause a decrease in metal removal during activated sludge treatment and as a consequence higher metal concentrations in the effluent (Rossin et al., 1983a). Metals whose concentration may increase in the

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R. PERRY et al.

effluent inclade aluminium, cadmium, chromium. copper, iron, lead, nickel and zinc (Stoveland et al.. 1979b: Wei et al., 1979: Rossin et al., 1983a). The NTA concentrations may also be high enough to cause mobilisation of metals from sediments (Miiller and F6rstner, 1976: Dietz. 1982). Indeed. estimated concentrations of NTA in the River Rhine were NTA to be introduced in detergents in Europe range from 0.04 to 0.7 mg 1-~ at the Dutch Border (Salomons and Pagee. 198l; Stehfest, 1982) and under some local conditions much higher concentrations can be expected (Salomons and Pagee, 1981). During a monitoring survey in Canada, Woodiwiss et al. (1979) report one case when the NTA concentration detected in a receiving stream below a sewage treatment plant outfall was as high as 2.46 mg 1-~. The elevated concentrations of metals either from sewage effluents or mobilised from sediments could result in the infringement of standards for raw waters (European Economic Community, 1975) and water for potable supply (World Health Organisation, 1971: European Economic Community, 1980) since conventional water treatment processes are not well suited to safeguard against the passage of trace metals (F6rstner and Whittmann, 1979). Nitrilotriacetic acid discharged to rivers during cold weather conditions will not biodegrade as rapidly even after acclimatisation compared to periods in high water temperatures (Larson et al., 1981). Thus NTA present in rivers and lakes could upon abstraction enter potable supplies. In the United Kingdom rivers where water reuse occurs account for 32~, of actual abstractions for public water supphes (Central Water Planning Unit, 1980). At points of major abstraction in the rivers Thames, Lea and Great Ouse sewage effluent can account for 50-60~o of the total flow (Fielding and Packham, 1977). In the Thames Water Authority area as a whole 58~/of the potable is abstracted from lowland rivers receiving sewage effluent (Central Water Planning Unit, 1980). The higher expected concentrations of NTA in raw sewage, lower dilution of sewage effluent by rivers and greater water reuse in Europe compared to Canada demonstrate the inadequacy of Canadian field studies and monitoring data when applied to the European situation. In addition, in the United Kingdom, as a result of the high population density, typical sewage treatment works serve a substantial population equivalent. Only the treatment works described in Bouveng et al. (1970) and Shumate et al. (1970) and some in Woodiwiss et al. (1979) were of this size. In Europe retention times are typically 3.5-7 h. Wei et al. (1979) studied a plant within this range but retention times (where reported) were higher in all other studies (Table 10). Variations in influent and effluent concentrations are subject to reduction due to composite sampling. Results published from field studies have been daily, weekly or monthly averages based upon composite samples ranging from 4 h to a week or discrete

samples (Tables 10). Thus no short term variations in effiuent NTA concentration, which may be high enough to cause mobilisation of heavy metals to receiving waters, are revealed. For example, if the effluent NTA concentration is 0.1 mg 1-~ for 90°0 of the time and 5 m g l -~ for 10go of the time the composite sample value would be 0.59 mg 1-~. This is apparently not significant and assuming an influent NTA concentration of 10mgl -~ would indicate a removal efficiency greater than 94°~o. However, during the 10°0 of the time that the effluent NTA concentration approached 5 mg I-* there would have been substantial metal mobilisation and serious contamination of the receiving waters with NTA as well as metals. It is important to recognise that only an intermittent t:ai[ure in NTA biodegradation is required to cause substantial contamination of the receiving water with heavy metals. Since biological wastewater treatment removes 70-80°'0 of the metals it receives (and the influent concentrations would be elevated in the presence of NTA because of the reduced removals during primary sedimentation) concentrations of metals in the biological solids are typically two orders of magnitude greater than the influent concentrations. The metals concentrated into the biological solids are retained in the biological treatment plant for a period equal to the mean cell residence time (sludge age). By definition, at a sludge age of 8 days 50°/ of the biomass will be retained within the treatment plant over this period together with the associated heavy metals removed during this period. Thus at any one time 50'/0 of the heavy metals removed during a period of 8 days can potentially be mobilised by NTA should any failure in its biodegradation occur. A failure in NTA biodegradation lasting only a few hours could be sufficient, depending upon the NTA concentration, to mobilise nearly all of these metals, causing substantial contamination of the receiving water. This leads to considerable concern about the effluent NTA concentration when it is reported by Bouveng et al. (1970) that removal based on composite samples was as low as 30Vo on some occasions. Even the most thorough study in Canada, by Wei et al. (1979), revealed NTA removals over 24 h composites as low as 58~o. The survey by Woodiwiss et al. (1979) does not Nve overall removals. Only monthly discrete samples were collected with no allowance for retention times through the plants and only geometric mean influent and effluent NTA concentrations and ranges were presented. The ranges in effluent NTA concentrations indicate, however, that there was considerable variation in NTA biodegradation, in several cases the highest recorded effluent NTA concentrations were greater than the geometric mean influent NTA concentrations. Field studies and monitoring programmes almost invariably undertaken at low influent NTA concentrations have not only been subject to inadequate sampling regimes but have also been hampered by

NTA as a detergent builder analytical difficulties. Interference reported by Games (1981) and Kirk et al. (1982b) using the gas-liquid chromatographic method of Aue et al. (1972) leads to the conclusion that the results of programmes using this technique, without adequate regard to interference, may be subject to significant errors. In addition, variations between laboratories are evident even when applying identical techniques. Modification of the gas-liquid chromatographic method to incorporate recovery determinations by NTA radiolabetling lengthens the analysis time, which is already appreciable, and introduces a technique which is not generally available nor suitable for routine analysis. Following a comprehensive comparison of gas-liquid chromatography, differential pulse polarography and a colorimetric method Kirk et al. (1982b) found differential pulse polarography to be the most precise technique down to concentrations of 0.100mgl -~ of NTA, and applicable down to 0.025 mgl -~ of NTA in non-saline matrices, whilst gas-liquid chromatography was only normally required at NTA concentrations below 0.025 mg l- t. In addition, differential pulse polarography is the only technique applicable to saline samples. NTA removal during the primary sedimentation of sewage has been concluded to be attributable to adsorption onto the sludge solids, thereby increasing the potential NTA load on sludge treatment and disposal practices. Removal by adsorption would appear to be insufficient to prevent a reduction in heavy metal removal during primary treatment. Current U.K. guidelines for the disposal of sewage sludge to agricultural land are however restrictive for raw sludges (Department of the Environment and National Water Council, 1981) and sludge stabilisation, primarily by anaerobic digestion, may become a mandatory requirement throughout the EEC (Commission of the European Communities, 1982). Therefore the behaviour of NTA during anaerobic digestion will, in part, determine the potential contamination of land and sea by the disposal of sewage sludge. The removal of NTA during anaerobic digestion has been shown to be dependent on the type of sludge being treated and therefore any biodegradation is subject to variation between works in terms of works design and the possible introduction of new treatment practices; separate digestion of primary (anaerobic digestion) and waste activated sludge (aerobic digestion) has been advocated (Grant et al., 1981). Septic tank treatment, which is widely used in rural areas in Europe and the United States, would result in groundwater contamination by NTA. This particularly important since use of groundwater as a domestic supply is of increasing importance, and in certain areas, septic tank drainfields constitute a significant groundwater input sometimes resulting in groundwater well pollution (Dewalle and Schaff, I980). Untreated and treated discharges of sewage and et al.

WR

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sewage sludge would contaminate estuarine and costal waters and could result in significant NTA concentrations since biodegradation of NTA in saline environments has been demonstrated to be minimal. Concern over such NTA recalcitrance has been specifically concentrated on the metal chelating ability of NTA, although Eisler et al. (1972) concluded that NTA p e r se might present a significant hazard to marine teleosts and higher invertebrates even when used only as a partial replacement for STPP in detergents. The discharge of sewage to coastal waters has a marked effect on metal concentrations in the discharge area (Martin et al., 1976) and metal mobility in estuarine and coastal waters may be enhanced by chelation with NTA (Niirnberg and Raspor, 1981 : Kirk et al., 1983a). The effect of such changes in metal speciation has been the subject of considerable debate. Whilst it has been suggested that the stronger the metal complex the lower the toxicity of a given concentration of total metal (Allen et al., 1980), there have also been reports of enhanced toxicity of chelated trace metals (Haberer and Norman, 1972: George and Coombs, 1977). Niirnberg (1980) reported the accumulation of chelates at biological interfaces, thereby increasing toxic metal concentrations at the site of entry to organisms, and potentially resulting in the transfer of metals along marine food chains back to man and the terrestrial environment. Conversely, it has been demonstrated that the growth of toxic "red tide" dinoflagellates can be promoted by the chelation of copper by NTA which results in a reduction in the toxic free ionic copper concentration which previously inhibited growth (Gran~li and Edler, 1983). Of the 67~ of sludge which is applied to land in the U.K. approximately half is applied to general arable and grazing land (Department of the Environment and National Water Council, 1981). Since only 13~ of the total flow of sewage to a works is produced as sludge, and this typically contains more than 50-80~,,; of the total quantity of heavy metals entering the works, the metals are concentrated to a significant degree. It is generally accepted that the impact of heavy metals in sludge applied to land will depend on their speciation and thus any interaction between NTA and heavy metals in sludge following disposal must be considered. Groundwater contamination by NTA and metal-NTA chelates may result from the disposal of contaminated sludge to land in addition to septic tank final effluent discharge to percolation fields. Reports of NTA removal in soil have been contradictory although it is generally concluded that biodegradation is insignificant in the saturated zone and that adsorption is sufficient to retard, but not prevent, movement to groundwater. The distribution and movement of NTA in agricultural soil is particularly influenced by the method of sludge application and subsequent soil treatment (Kirk et al.. 1983b) and this appears likely to influence significantly the ability of NTA to enter groundwaters. As a powerful

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R. PERRY et al.

metal chelating agent N T A may increase the p l a n t uptake of heavy metals, particularly in calcareous soils (Wallace et al.. 1974) a n d the potential consequences of such effects must be viewed with contern. REFERENCES Abdulla I. and Smith M. S. (1963) Influence of chelating agents on the concentration of some nutrients for plants growing under acid and under alkaline conditions. J. Sci. Fd Agric. 14, 98-109. Afghan B. K. and Goulden P. D. (1971) Determination of trace quantities of nitrilotriacetic acid by differential cathode-ray polarography. Enrir. Sci. Technol. 5, 601-606. Allen H. E. and Boonlayangoor C. (t978) Mobilization of metals from sediments by NTA. Verh. int. Verein. Limnol. 20, 1956-1962. Allen H. E. and Unger M. T. (1980) Evaluation of potential metal mobilization from aquatic sediments by comptexing agents. Z. Wasser, Abwasser Forsch. 13, 124-129. Allen H. E., Hall R. H. and Brisbin T. D. (1980) Metal speciation. Effects on aquatic toxicity. Ent'ir. Sci. Technol. 14, 441-443. Anderson D. M. and Morel F. M. M. (1978) Copper sensitivity of Gonyaulax tamarensis. LimnoL Oceanogr. 23, 283-295. Appleby D. J. (1977) The impact of phosphorus control activities. The experience in Ontario. Erco Industries Ltd, Toronto. Ontario. Aue W. A., Hasting C. R., Gerhardt K. O., Pierce J. O., Hill H. H. and Mosemann R. F. (1972) The determination of part-per-billion levels of citric and nitrilotriacetic acids in tap water and sewage effluents. J. Chromat. 72, 259-267. Baalsrud K. (1982)The rehabilitation of Norway's largest lake. Water Sci. Technol. 14, 21-30. Balmer P. (1982) Water pollution control in Sweden. Water Qual. Bull. 7, 127-130, 151-[52. Banat K., Frrstner U. and Miiller G. (1974) Experimental mobilization of metals from aquatic sediments by nitrilotriacetic acid. Chem. Geol. 14, 199-207. Barth E. F.. Tabak H. H. and Mashi C. I. (1979) Biodegradation studies of carboxymethy!tartronate, U.S. Environmental Protection Agency, Report EPA-600/2-78115, 35 pp. Bartholomew G. W. and Pfaender F. K. (1983) Influence of spatial and temporal variations on organic pollutant biodegradation rates in an estuarine environment. Appl. envir. Microbiol. 45, 103-109. Beccari M. (1982) Water pollution control technology in Italy. Water Qual. Bull. 7, 135-140. Berth P. (1978) Recent development in the field of inorganic builders. J. Am. Oil Chem. Soc. 55, 52-57. Birkner L. (1982) Detergent--environmental considerations. Second Symposium on Technological, Environmental and Economic Trends in Detergency. Rome lnstituto Di Merceologia Dell'Universita Di Roma. BjSrndal H., Bouveng H. O,, Solyom P. and Werner J. (1972) NTA in sewage treatment. Part 3. Biochemical stability of some metal chelates. Vatten 28, 5-16. Bott T., Patrick R., Larson R, and Rhyne C. (1980) The effect of nitrilotriacetic acid (NTA) on the structure and functioning of aquatic communities in streams. U.S. Environmental Protection Agency, Report No. EPA600/3-80-050. Bourquin A. W. and Pr-zybyszewski V. A. (1977) Distribution of bacteria with nitrilotriacetate degrading potential in an estuarine environment. Appl. enrir. Microbiol. 34, 411-418. Bouveng H. O., Davisson G. and Steinberg E. (1968) NTA in sewage treatment. Vatten 24, 348-359.

Bouveng H. O., Solyom P. and Werner J. (1970) NTA in sewage treatment. Part 2. Degradation of NTA in a filter and oxidation pond. Vatten 26, 389-398. Brown M. J. and Lester J. N. (1979) Metal removal in activated sludge: the role of bacterial extracellular polymers. Water Res. 13, 817-837. Bunch R. L. (1982) Water pollution control technology m the U.S.A. Water Qual. Bull. 7, 107-112. 149-150. Bundesgesundheitsamt (1982) NTA in Waschmitteln. Communication to. Bundesminister ffir Jugend. Familie und Gesundheit, Berlin. Carrondo M. J. T., Perry R. and Lester J N t1980) Behaviour of zeolite type A in the activated sludge process--l. Influence on treatment parameters. Y. W~tt. Pollut. Control Fed. 52, 2796-2800. Carrondo M. J. T., Perry R. and Lester J. N. (1981a) Sedimentation of zeolite type A in water and waste water. Can. J. cit,. Engng 8, 206-217. Carrondo M. J. T., Perry R. and Lester J. N. (1981b). Type A zeolite in the activated sludge process--II. Heavy metal removal. J. Wat. Pollut. Control Fed. 53, 344--35t. Central Water Planning Unit (1980) Re-use of water for potable supplies. Project 4, Final Report, 1980. Reading, U.K. Chau Y. K. and Shiomi M. T. (1972) Comptexing properties of nitrilotriacetic acid in the Lake environment. War. Air Soil Pollut. 1, 149-164. Cheng M. H., Patterson J. W. and Minear R. A. (1975) Heavy metals uptake by activated sludge. J. Wat. Pollut. Control Fed. 47, 362-376. Cleasby J, L., Hubly D. W., Ladd T. A. and Schon E. A. (t974) Treatment of waste containing NTA in a trickling filter. J. Wat. Pollut. Control Fed. 46, 1973-!987. Claesson A. (1971) Anaerobic bacterial degradation of nitrilotriacetate NTA. Vatten 27, 410-411. Collinge V. K. and Bruce A. M. (1981) Sewage sludge disposal: a strategic review and assessment of research needs. Water Research Centre Technical Report, TR 166, 31 pp. Commission of the European Communities (1982) Proposal for a council directive on the use of sewage sludge in agriculture. Off. J. Eur. Commun. C264/3-C264/7, Constantin M. J. and Owens E. T. (1982) Introduction and perspectives of plant genetic and cytogenetic assays. A report of the U.S. Environmental Protection Agency Gene-Tox Program. Mut. Res, 99, 1-12. Department of the Environment (1971) Twelfth Progress Report o f the Standing Technical Committee on Synthetic Detergents. HMSO, London. Department of the Environment (1978) Eighteenth Progress Report o f the Standing Technical Committee orr Synthetic Detergents. HMSO, London. Department of the Environment (!980) Twentieth and final report of the Standing Technical Committee on Synthetic Detergents. HMSO, London. Department of the Environment and National Water Council (1981) Report of the Sub-Committee on the Disposal of Sewage Sludge to Land, Standing Technical Committee Report No. 20, National Water Council, London. DeWalle F. B. (1981) Failure analysis of large septic tank systems. J. envir. Engng Div. Am. Soc. cit'. Engrs 107, EE 1, 229-240. DeWalle F. B. and Schaff R. M. (1980) Ground water pollution by septic tank drainfields. J. era'iron. Engng Dit,. Am. Soc. cir. Engrs 106, EE3, 631-648. Dietz F. (1982) Zur frage der remobilisierung von schwermetallen durch nitrilotriessigs~.ure (NTA). Korresp. Abwasser 29, 692-693. Dunlap W. J., Crosby R. L., McNabb J. F.. Btedsoe B. E. and Scalf M. R. ( 1971) Investigation concerning probable impact of nitrilotriacetic acid on ground water: Environmental Protection Agency, Water Pollution Research Series 16060 GHR.

NTA as a detergent builder Dunlap W. J., Crosby R. L., McNabb J. F.. Bledsoe B. E. and Scalf M. R. (1972) Probable impact of NTA on ground water. Ground Water 10, 107-117. Eden G. E., Culley G. E. and Rootham R. C. ([972) Effect of temperature on the removal of NTA (nitrilotriacetic acid) during sewage treatment. Water Res. 6, 877-883. Eisler R., Gardner G. R.. Hennekey R. J.. LaRoche G., Walsh D. F. and Yevich P. P. (1972)Acute toxicology of sodium nitrilotriacetic acid (NTA) and NTA-containing detergents to marine organisms. Water Res. 6, 1009-1027. Elliott H. A. and Denneny C. M. (1982) Soil adsorption of cadmium from solutions containing organic ligands. J. era'iron. Qual. 11, 658-663. Elliott H. A. and Huang C. P. (1979) The adsorption characteristics of Cu(II) in the presence of chelating agents. J. Colloid Interface Sci. 70, 29-45. Enfors S. O. and Molin N. (1971) Anaerobic degradation of nitrilotriacetate (NTA) by bacteria. Vatten 27, 162-163. Enfors S. O. and Molin N. (1973a) Biodegradation of nitrilotriacetate (NTA) by bacteria. I. Isolation of bacteria able to grow anaerobically with NTA as a sole carbon source. Water Res. 7, 881-888. Enfors S. O. and Mo[in N. (1973b) Biodegradation of nitrilotriacetate (NTA) by bacteria. II. Cultivation of a NTA-degrading bacterium in anaerobic medium. Water Res. 7, 889-893. Environmental Protection Agency (1980) Final Report: NTA. (NTA Risk Assessment). Management Support Division, Office of Pesticides and Toxic Substances, Washington, DC. Epstein S. S. (1972) Toxicological and environmental implications on the use of nitrilotriacetic acid as a detergent builder--ll. Int. J. ent'ir. Stud. 3, 13-21. Erickson S. J., Maloney T. E. and Gentile J. H. (1970) Effect of nitrilotriacetic acid on the growth and metabolism of estuarine phytoplankton. J. Wat. Pollut. Control Fed. 42, R329-335. European Economic Community (1975) Council directive concerning the quality of water intended for the abstraction of drinking water in the member states (75/440/EEC). Off~ J. Eur. Commun. LI94/26-LI94/31. European Economic Community (1980) Council directive relating to the quality of water intended for human consumption (80/778/EEC). Off. J. Eur. Commun. L229/I 1- L229/29. Federal Register (1980) Nitrilotriacetic acid: denial of citizens' petition to initiate regulatory proceedings prohibiting the manufacture and distribution. Fed. Reg, 45, 72778-72780. Feliciano D. V. (1982) Fact sheet for wastewater treatment. J. Wat. Pollut. Control Fed. 54, 1346-1348. Fielding M. and Packham R. F. (1977) Organic compounds in drinking water and public health. J. Inst. War. Engrs Sci. 31, 353-375. Fish H. (1983) Sea disposal of sludges: the U.K. experience. Water Sci. Technol. 15, 77-88. Fleckseder H. (1982) Water pollution control in Austria. Water Qual. Bull. 7, 145-149. Flynn K. C. (1982) New challenges in the Great Lake States to banning phosphorus in detergents. J. War. Pollut. Control Feet. 54, 1342-1345. Forsberg C. and Lindquist G. (1967) Experimental studies on bacterial degradation of nitritotriacetate. Vatten 23, 265-277. F6rstner U. and Wittmann G. T. W. (1979) Metal Pollution in the Aquatic Environment. Springer-Verlag, New York. ~Gakstatter J. H., Bartsch A. F. and Callahan C. A. (1978) The impact of broadly applied effluent phosphorus standards on eutrophication control. Water Resour. Res. 14, 1155-1158.

Games L. M., Staubaeh J. A. and Kappeler T. U. (1981) Analysis of nitrilotriacetic acid in environmental waters. Tenside Deterg. 18, 262-265.

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