Environmental Fate and Microbial Degradation of Aminopolycarboxylic Acids

FEMS Microbiology Reviews 25 (2001) 69^106 www.fems-microbiology.org

Environmental fate and microbial degradation of aminopolycarboxylic acids

Margarete Bucheli-Witschel 1, Thomas Egli *

Swiss Federal Institute for Environmental Science and Technology, Department of Microbiology, Uë berlandstrasse 133, CH-8600 Du«bendorf, Switzerland

Received 18 January 2000; received in revised form 30 August 2000; accepted 30 August 2000

Abstract

Aminopolycarboxylic acids (APCAs) have the ability to form stable, water-soluble complexes with di- and trivalent metal ions. For that reason, synthetic APCAs are used in a broad range of domestic products and industrial applications to control solubility and precipitation of metal ions. Because most of these applications are water-based, APCAs are disposed of in wastewater and reach thus sewage treatment plants and the environment, where they undergo abiotic and/or biotic degradation processes. Recently, also natural APCAs have been described which are produced by plants or micro-organisms and are involved in the metal uptake by these organisms. For the two most widely used APCAs, nitrilotriacetate (NTA) and ethylenediaminetetraacetate (EDTA), transformation and mineralisation processes have been studied rather well, while for other xenobiotic APCAs and for the naturally occurring APCAs little is known on their fate in the environment. Whereas NTA is mainly degraded by bacteria under both oxic and anoxic conditions, biodegradation is apparently of minor importance for the environmental fate of EDTA. Photodegradation of iron(III)-complexed EDTA is supposed to be mostly responsible for its elimination. Isolation of a number of NTA- and EDTA-utilising bacterial strains has been reported and the spectrum of APCAs utilised by the different isolates indicates that some of them are able to utilise a range of different APCAs whereas others seem to be restricted to one compound. The two best characterised obligately aerobic NTA-utilising genera (Chelatobacter and Chelatococcus) are members of the K-subgroup of Proteobacteria. There is good evidence that they are present in fairly high numbers in surface waters, soils and sewage treatment plants. The key enzymes involved in NTA degradation in Chelatobacter and Chelatococcus have been isolated and characterised. The two first catabolic steps are catalysed by a monooxygenase (NTA MO) and a membrane-bound iminodiacetate dehydrogenase. NTA MO has been cloned and sequenced and its regulation as a function of growth conditions has been studied. Under denitrifying conditions, NTA catabolism is catalysed by a NTA dehydrogenase. EDTA breakdown was found to be initiated by a MO also which shares many characteristics with NTA MO from strictly aerobic NTA-degrading bacteria. In contrast, degradation of [S,S]-ethylenediaminedisuccinate ([S,S]-EDDS), a structural isomer of EDTA, was shown to be catalysed by an EDDS lyase in both an EDTA degrader and in a NTA- utilising Chelatococcus strain. So far, transport of APCAs into cells has only been studied for EDTA and the results obtained give strong evidence for an energy-dependent carrier system and Ca2 seems to be co-transported with EDTA. Due to their metal-complexing capacities, APCAs occur in the environment mostly in the metal-complexed form. Hence, the influence of metal speciation on various degradation processes is of utmost importance to understand the environmental behaviour of these compounds. In case of biodegradation, the effect of metal speciation is rather difficult to assess at the whole cell level and therefore only limited good data are available. In contrast, the influence of metal speciation on the intracellular enzymatic breakdown of APCAs is rather well documented but no generalising pattern applicable to all enzymes was found. ß 2001 Federation of European Microbiological Societies. Published by Elsevier Science B.V. All rights reserved.

Keywords: Abiotic elimination mechanism; Aminopolycarboxylic acid; Biochemistry; Biodegradation; Chelation; Ethylenediaminedisuccinate; Ethylenediaminetetraacetate; Metal speciation of aminopolycarboxylic acid; Metal-complexing agent; Nitrilotriacetate

* Corresponding author. Tel.: +41 (1) 823-51-58; Fax: +21 (1) 823-55-47; E-mail: [email protected]

1 Present address: Department of Microbiology, Stockholm University, S-106 91 Stockholm, Sweden.

Abbreviations: L-ADA, L-alaninediacetate; AEAA, N-(2-aminoethyl)aspartic acid; APCA, aminopolycarboxylic acid; ASDA, asparaginic acid diacetate; DH, dehydrogenase; DTPA, diethylenetriaminepentaacetate; ED3A, ethylenediaminetriacetate; EDDA, ethylenediaminediacetate (and its isomers N,N- EDDA and N,NP-EDDA); EDDS, ethylenediaminedisuccinate; EDTA, ethylenediaminetetraacetate; EDMA, ethylenediaminemonoacetate; HEDTA, hydroxyethylethylenediaminetriacetate; IDA, iminodiacetate; MGDA, methylglycinediacetate; MO, monooxygenase; NTA, nitrilotriacetate; NtrR, nitrate reductase; PDTA, 1,3-propylenediaminetetraacetate; PMS, phenazinemethosulfate; SDA, serinediacetate

0168-6445 / 01 / $20.00 ß 2001 Federation of European Microbiological Societies. Published by Elsevier Science B.V. All rights reserved. PII: S0168-6445(00)00055-3 70 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Contents

1. Introduction ...... 70 1.1. Aminopolycarboxylic acids (APCAs) ^ an important group of chelating agents . . . . . 70 1.2. Important APCAs ...... 73 1.3. Fields of application for APCAs ...... 76 1.4. Concentrations of APCAs in the environment ...... 77 1.5. Speciation of APCAs in the environment ...... 78 1.6. Environmental risks ...... 79 2. Abiotic elimination of APCAs in the environment ...... 80 2.1. NTA ...... 80 2.2. EDTA ...... 80 2.3. Other APCAs ...... 81 3. Microbial degradation of APCAs during wastewater treatment and in the environment . . . 82 3.1. NTA ...... 82 3.2. EDTA ...... 83 3.3. DTPA ...... 84 3.4. Other synthetic APCAs ...... 84 3.5. Naturally occurring APCAs ...... 84 4. Degradation of APCAs by pure microbial cultures ...... 84 4.1. Pure cultures utilising NTA ...... 85 4.2. Pure cultures utilising EDTA ...... 86 4.3. Potential of bacterial isolates for the treatment of APCA-containing wastewaters . . . . 87 5. Biochemistry and genetics of APCA degradation ...... 87 5.1. NTA degradation ...... 88 5.2. Catabolism of EDTA and the similarity of EDTA MO with NTA MO ...... 91 5.3. EDDS catabolism ...... 93 6. Regulation of the APCA degradation and enzyme expression ...... 93 6.1. Regulation of NTA degradation ...... 93 6.2. Regulation of degradation of EDTA and EDDS ...... 95 7. In£uence of metal speciation on microbial APCA degradation ...... 95 7.1. In whole cells ...... 95 7.2. At the enzyme level ...... 97 8. Outlook ...... 97

Acknowledgements ...... 99

References ...... 99

1. Introduction Additionally, the presence of basic secondary or tertiary amino groups and the large negative charge of APCAs 1.1. Aminopolycarboxylic acids (APCAs) ^ an important contributes to the high stability of metal^APCA com- group of chelating agents plexes. To describe the stability of such a metal^ligand complex, the equilibrium constant K is used. It is de¢ned APCAs, also known as complexones, are compounds as that contain several carboxylate groups bound to one or MeL more nitrogen atoms (Fig. 1). Hence, they are essentially K 1 MeWL derived from the amino acid glycine [1,2]. The major chemical property of APCAs is their ability where [Me] is the concentration of the metal ion, [L] the to form stable and water-soluble complexes with many concentration of the ligand, and [MeL] that of the ligand^ metal ions. APCAs form a special kind of metal complexes metal complex at equilibrium [2]. because they co-ordinate the metal ion by forming one or In many industrial processes and products, the presence more heteroatomic rings (Fig. 2). This ring formation of free metal ions causes problems such as the formation leads to a higher stability of the complexes as compared of insoluble metal salt precipitates or the catalysis of the to metal^ligand complexes in which no such rings are decomposition of organic compounds. Chelation masks present, a phenomenon which is called `chelate e¡ect'. the metal ions and restricts them from playing their nor- M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 71 mal chemical role and from entering potentially harmful structure is shown in Fig. 1A. In recent years, an increas- and unwanted reactions. Because of their industrial impor- ing number of APCAs of biological origin have been de- tance, chelating agents are produced and used in large scribed, most of which are involved in metal acquisition by quantities. Among the chelating agents presently em- organisms (Fig. 1B). ployed, polyphosphates are the most widely used compounds, followed by APCAs [3]. Therefore, much atten- 1.2.1. Synthetically produced APCAs tion has been paid to the environmental fate of APCAs, In 1862, the ¢rst chemical synthesis of an APCA, with interest focused on nitrilotriacetate (NTA) and ethyl- namely that of NTA, was described [8]. It was based on enediaminetetraacetate (EDTA), the two most important the reaction of monochloroacetic acid in an ammonical APCAs. In the case of NTA, research concentrated on solution. Much later, in 1935, the synthesis of EDTA biodegradation because this was found to be the major was reported by I.G. Farbenindustrie. Again, the synthesis environmental elimination pathway for this compound. was based on monochloroacetic acid reacting with ethyl- Several reviews have already been dedicated to this topic enediamine in the presence of sodium hydroxide. Alterna- [4^6]. For EDTA, however, biodegradation does not tively, EDTA is produced from ethylenediamine reacting seem to be a relevant elimination mechanism and so far with sodium cyanide and formaldehyde in the presence of an abiotic mechanism has been identi¢ed as major process sodium hydroxide. Depending on the amine employed, for the degradation of EDTA in natural systems [7]. also other APCAs, for instance 1,3-propylenediaminete- However, an increasing number of reports on the biode- traacetate (PDTA), can be produced by this type of reac- gradation of EDTA have been published recently and also tion [2]. other APCAs are receiving more and more attention con- The most important function of APCAs is to form met- cerning their occurrence and behaviour in the environ- al chelates. This chelation has a strong in£uence on their ment. environmental fate, including biodegradation. We will In this review, we will summarise the information con- therefore summarise the properties of the most intensively cerning the microbial breakdown of APCAs. At the same investigated APCA chelates, i.e. the metal complexes of time, we would like to present the most important aspects NTA and EDTA. of the abiotic degradation of this group of compounds to NTA contains four donor atoms and is a so-called provide a current summary of the knowledge of the fate of quadridentate chelating ligand. It forms 1:1 complexes APCAs in the environment. with metal ions by establishing three chelate rings with four co-ordination sites of the metal occupied by NTA. 1.2. Important APCAs Since most cations have a co-ordination number of six, the remaining two sites are normally occupied by water mol- The predominant representatives of APCAs are the syn- ecules, resulting in octahedral co-ordination of the ion thetically produced compounds EDTA, NTA, diethylene- (Fig. 2A) [1]. triaminepentaacetate (DTPA) and hydroxyethylethyl- EDTA contains six donor atoms and acts as a hexaden- enediaminetriacetate (HEDTA). Figures on the total tate ligand. It can form a maximum of ¢ve chelate rings. amounts produced of the most important synthetic Ideally, EDTA should form an octahedral complex in APCAs are di¤cult to obtain and the estimated amounts which both the metal ion and EDTA have a co-ordination of these chelating agents produced or used in the USA and number of six (Fig. 2A). However, this octahedral co-or- Western Europe are listed in Table 1. Their chemical dination seems to be only possible with cations of rela- tively small size. With larger cations, constraints within the structure of the EDTA ligand prevent this ideal structure, and the complexed metal ion may still remain acces- Table 1 Estimated use of synthetic chelating APCAs (in 103 metric tons) in the sible to other ligands such as water molecules. Indeed, X- USA and Western Europe in 1981 ray analysis has shown that the structure of most metal USA Western Europe EDTA complexes di¡ers from the ideal octahedral structure and that the cations exhibit often higher co-ordina- EDTA 42 13.6 NTA 32 8.3 tion numbers than six (Fig. 2B). On the other hand, in 2 2 DTPA 4 0.5 some complexes, such as those with Cu or Ni , EDTA HEDTA 18 2.0 does not fully utilise its donor capacity by leaving one Other APCAs 2.5 2.0 carboxylate group not co-ordinated. Octahedral co-ordi- Sodiumpolyphosphates 586 1111 nation is instead completed by a water molecule (Fig. 2B) Organophosphonates 10 10 Hydroxycarboxylic acids 110 15 [1,2]. Stability constants for 1:1 complexes of EDTA and For comparison, data are also shown for other industrially important complexing agents. Figures for the USA are based on production, those NTA are listed in Table 2. For each metal, the stability for Europe are estimated from information on consumption (adapted of its NTA complex is several orders of magnitude lower from [3]). than that of its EDTA complex, as can be expected from 72 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Fig. 1. Structural formulas of some of the important synthetically produced (A) and naturally occurring (B) APCAs. the lower chelating capability of NTA. Although DTPA or HEDTA and it does not decompose even in Fe(III)EDTA has a stability constant of log K = 25.0, the most strongly alkaline solutions [2]. this is still not high enough to keep it from decomposing Lately, the synthesis of several new APCAs was re- at pH values above 8^9 due to precipitation of iron(III)- ported, namely L-alaninediacetate (L-ADA), serinediace- hydroxide. In the search for more e¡ective chelating tate (SDA), asparaginic acid diacetate (ASDA) and meth- agents at higher pH values, DTPA and HEDTA were ylglycinediacetate (MGDA) (Fig. 1). They are derived developed. They are preferentially employed instead of from amino acids by substituting the amino group with EDTA for sequestering iron(III) ions in the pH range of two acetyl groups. The stability of their metal complexes 8^10. EDDHA with a log K of 33 (Fig. 1A) is able to seems to be closer to NTA complexes than to EDTA^ complex iron(III) ions more selectively than EDTA, metal complexes [9]. Also, the synthesis of N-acyl deriva- M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 73

Fig. 1 (continued).

tives of ethylenediaminetriacetate (ED3A), chemicals 1.2.2. Naturally occurring APCAs that act both as a surfactant and powerful chelating Recently, it was shown that APCAs are also naturally agent, given the short name chelactant, has been described occurring compounds. Some even contain an ethylenedi- [10]. amine central moiety, a feature that was considered to 74 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Fig. 2. Ideal octahedral structure of metal^NTA and metal^EDTA complexes (A) as well as steric structure of NiEDTA and Fe(III)EDTA as determined by X-ray analysis of solid complexes [255,256] (B). be restricted to xenobiotic chelating agents such as EDTA. Table 2 1.2.2.1. Ethylenediaminedisuccinate (EDDS). The ¢rst Stability constants of 1:1 complexes of NTA, EDTA and [S,S]-EDDS with di- and trivalent metal ions determined for an ionic strength of 0.1 natural APCA described, EDDS, was isolated from cul- M at a temperature of 25³C or ^ where indicated (+) ^ at 20³C (¢gures ture ¢ltrate of the actinomycete Amycolatopsis orientalis.It were taken from [254]) was detected in an antibiotic screening program due to its Metal ion Log KMeNTA Log KMeEDTA Log KMeEDDS ability to inhibit activity of the Zn2-dependent phospho- Mg2 5.47 8.83 5.82 lipase C [11]. In the actinomycete, EDDS is most probably Ca2 6.39 10.61 4.23 2 involved in Zn uptake, because its production was Mn2 7.46 13.81 8.95 (+) found to be totally repressed at Zn2 concentrations high- Zn2 10.66 16.44 13.49 (+) er than 2.5 WM in the growth medium [12], whereas below Co2 10.38 16.26 14.06 2 2 Cu 12.94 18.7 18.36 2.5 WM it increased with decreasing Zn concentrations. 2 2 3 2 2 Pb 11.34 17.88 12.7 (+) In contrast, other trace metals (Fe ,Fe ,Co ,Mn , Cd2 9.78 16.36 10.8 (+) 2 Mo ) exerted no repressive e¡ect on EDDS production Al3 11.4 16.5 [13]. However, there is no preferential complexation of Fe2 8.33 (+) 14.27 Zn2 by EDDS, for the stability constants for Fe3, Fe3 15.9 25.0 22.0 (+) 2 Co2,Ni2 and Cu2 are higher than that for Zn2 (Table Ni 11.5 18.52 16.79 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 75

2) [14]. Also, it has not been demonstrated yet that Zn2 is amid bonds to a 1,4-diaminobutane (putrescine) molecule really transported by EDDS into cells of A. orientalis. [23]. The asymmetric carbon atoms of the citric acid res- EDDS exhibits two chiral C-atoms resulting in the ex- idues are always in R,R-con¢guration. A value of 1023:5 istence of three di¡erent stereoisomers, [S,S]-, [R,S]- and was reported for the stability constant of the iron(III) [R,R]-EDDS. A. orientalis exclusively produces the S,S- rhizoferrin complex [24]. Thus, rhizoferrin is not a partic- isomer. For the biosynthesis of [S,S]-EDDS, two pathways ular powerful iron chelator at pH 7.4 compared to other were proposed by Cebulla [12], starting from either L-as- fungal and bacterial siderophores of the hydroxamate or partate and serine, or from oxaloacetate and 2,3-diamino- catecholate type but it becomes much more competitive at propionic acid. In both cases, N-(2-aminoethyl) aspartic pH values near 4^5, the growth optimum for the produc- acid (AEAA) would be an intermediate reacting then ing fungi [24]. with oxaloacetate to form [S,S]-EDDS. Transport experiments with 55Fe-labelled rhizoferrin EDTA is rather recalcitrant towards biological and con¢rmed that the compound acted as siderophore [21]. chemical degradation in the environment and there is in- In the producing fungal strain, Absidia spinosa, reductively creasing pressure to replace it by other powerful complex- inert gallium rhizoferrin, and the kinetically inert chromi- ing agents. For this, EDDS seems to be a promising can- um and rhodium complexes were also taken up, indicating didate. Consequently, its chemical and biotechnological that the intact metal^siderophore complex was transported production has been investigated. Two routes exist to syn- into the cell [25]. thesise EDDS chemically: ¢rstly, the reaction of maleic Rhizoferrin is able to deliver iron not only to the pro- anhydride with ethylenediamine yielding a mixture of the ducing organisms but to other micro-organisms as well. It three stereoisomers of EDDS [15,16], and secondly, the showed strong siderophore activity in nearly all bacterial reaction of aspartic acid with 1,2-dibromoethane leading strains tested from the Proteus^Providencia^Morganella to the formation of either pure [S,S]-EDDS or [R,R]- group, although these strains possess their own e¤cient EDDS, depending on the stereoisomer of the aspartic iron uptake system based on keto- or hydroxycarboxylate acid employed [17]. As an alternative, both the fermenta- iron complexes [26]. The rates of iron uptake in Morga- tion conditions for the production of [S,S]-EDDS by A. nella morganii mediated by rhizoferrin were equivalent to orientalis and the subsequent puri¢cation procedure were those measured in the presence of some of its own side- optimised [18]. Highest [S,S]-EDDS productivity was ob- rophores. Recently, a rhizoferrin receptor from M. morga- tained by fed-batch cultivation with a medium containing nii has been cloned [27]. This transport system (called very low Zn2 concentrations, glycerol as carbon source, a Rum, for rhizoferrin uptake into Morganella spp.) was mixture of glutamate and urea as nitrogen source, and speci¢c for ferric rhizoferrin and staphyloferrin A (de- high initial phosphate concentrations. Under optimised scribed below), while citrate and the citrate derivative conditions, ¢nal EDDS concentrations of some 20 g l31 aerobactin were not transported. The system consisted of were obtained. EDDS can be puri¢ed using a three-step two proteins, RumA located in the outer membrane and procedure consisting of an acid precipitation, an ethanol the periplasmic RumB. Homology analyses indicated a washing step and a ¢nal crystallisation leading to 92% close relationship of RumA to the outer membrane recep- purity with an overall yield close to 50% [18]. tor FecA of the Escherichia coli ferric citrate transport system. RumB, however, showed little homology to other 1.2.2.2. Rhizobactin. In 1984, Smith and Neiland [19] binding proteins with the exception of certain signatures of described the formation of the APCA rhizobactin by Rhi- conserved amino acid residues characteristic of binding zobium meliloti strain DM4 when cultured on a low-iron proteins of the siderophore transport family. medium. Consequently, rhizobactin is considered a sider- The biotechnological production of rhizoferrin and ophore. As EDDS, rhizobactin contains an ethylenedi- some of its analogues was also studied [28]. From several amine group with one amine function linked to pyruvic producing strains tested, Cunninghamella elegans and Mu- acid, and the second to the carbon skeleton of lysine car- cor rouxii exhibited highest productivity and were selected rying malic acid on the N6 amino group [20]. The absolute for process optimisation. Moreover, C. elegans was used ala lys malate con¢guration of rhizobactin was D , L , L , suggest- to produce analogues of rhizoferrin by feeding analogues ing that rhizobactin is biochemically related to the pyruvic of the two building blocks (1,4-diaminobutane and citric ala acid-derived opines which have a con¢guration of D , acid). Feeding citric acid analogues resulted in only low amino acid L . amounts of derivatives being formed. However, feeding of 1,5-diaminopentane or 1,3-diaminopropane caused a con- 1.2.2.3. Rhizoferrin. Also among fungal siderophores, siderable formation of analogues containing bridges with a representative of APCAs, rhizoferrin, was detected. Rhi- ¢ve or three methylene residues, respectively. Also, bis-(2- zoferrin was ¢rst isolated from cultures of Rhizopus micro- aminoethyl)-ether and branched diamine precursors were sporus var. rhizopodiformis [21]. Apparently, rhizoferrin is used as building blocks but only low product yields were a widespread siderophore within the class of Zygomycetes obtained. Finally, two interesting e¡ects were observed [22]. It is composed of two citric acid groups linked via when feeding the ketone 1,4-diamino-2-butanone as a pre- 76 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 cursor. Firstly, the synthesis of rhizoferrin itself was dra- logues, as well as avenic acid (Fig. 1B). These phytosider- matically reduced; and secondly, the analogue 2-oxorhizo- ophores are apparently widespread among grass species ferrin was formed preferentially, which was never observed and cultivars [33]. Mugineic acids might also be produced when other 1,4-diaminobutane analogues were supplied. and excreted by plants stressed by zinc de¢ciency [34]. This probably results from 1,4-diamino-2-butanone being Mass production of mugineic acid and its derivatives is a competitive inhibitor of ornithine decarboxylase, an en- proposed because of their potential application in agricul- zyme known to be responsible for the synthesis of 1,4- ture, medical science and pharmacy. Possible production diaminobutane in fungi, bacteria and plants. Hence, in methods take advantage of cell cultures of barley or C. elegans, diaminobutane seems to be a precursor of rhi- wheat, or are based on water cultures of intact plants, in zoferrin and ornithine decarboxylase is probably respon- both cases exposing the cells or plants to iron de¢cient sible for diaminobutane synthesis. Furthermore, feeding L- conditions [35]. ornithine led to a preferential excretion of rhizoferrin whereas addition of D-ornithine resulted in a signi¢cant 1.3. Fields of application for APCAs reduction of rhizoferrin produced. All these biologically produced rhizoferrin analogues had siderophore activity Due to their metal sequestering capacity, synthetic in growth promotion assays with M. morganii with activ- APCAs are used in many industrial processes or products ities comparable to or only slightly lower than that of in order to (i) prevent the formation of metal precipitates, rhizoferrin. In contrast, the citric acid analogues exhibited (ii) to hinder metal ion catalysis of unwanted chemical signi¢cantly lower activity. These results show clearly that reactions, (iii) to remove metal ions from systems, or (iv) also the structure of biologically produced APCAs can be to make metal ions more available by keeping them in manipulated, a point which might become interesting with solution. respect to a possible future exploitation of naturally occurring APCAs. 1.3.1. Use of APCAs to prevent formation of precipitates In addition to active washing ingredients, detergents 1.2.2.4. Staphyloferrins. Structurally related to rhizo- contain a large proportion of metal-complexing agents to ferrin is staphyloferrin A, a siderophore originally isolated inhibit the formation of insoluble Ca2 and Mg2 salts, from a culture of Staphylococcus hyicus [29,30]. It consists and thus prevent the deposition of scale on both textile of two citric acid residues linked by two amid bonds to D- ¢bres and washing machine parts. The ¢rst complexing ornithine [30]. As in the case of rhizoferrin, the tying to- agents employed in modern detergents were di- and tri- gether of two citric acid residues provides an iron-binding phosphates [36]. However, it was soon found that these agent much more powerful than citrate alone [29]. Staph- contribute to the eutrophication of lakes and rivers. Dur- ylococci produce also another siderophore of the complexing the search for substitutes, NTA was proposed as an one type, staphyloferrin B, the chemical structure of which alternative detergent builder and NTA-containing deter- is quite di¡erent from that of staphyloferrin A. Composed gents were ¢rst marketed in Sweden in 1967. Since then, of 2,3-diaminopropionic acid, citrate, ethylenediamine and NTA-based detergents have been used in Canada, Finland 2-ketoglutaric acid as structural components, it lacks the and Sweden, the US, and Switzerland, although they have symmetry of staphyloferrin A and rhizoferrin, and it is met considerable opposition [4] (note that EDTA is not therefore a less potent chelator of iron [26]. Both staph- used as a substitute for polyphosphates in household de- yloferrin siderophores are produced by a wide range of tergents, see below). In industrial cleaning agents, NTA, Staphylococcus strains, some of which were even found ETDA, and recently also MGDA, are used to prevent to produce both types [31]. Transport measurements in precipitation of calcium, magnesium and heavy metal salts S. hyicus with ferric staphyloferrin A and B revealed up- [37^39]. For the same purpose, EDTA is especially used in take of 55Fe within the range of 1^15 min with staphylo- the textile and photographic industry, as well as in electro- ferrin B being less e¡ective than staphyloferrin A [26]. plating processes instead of cyanide [37]. In the photographic industry, FeNH4EDTA is also employed as oxi- 1.2.2.5. Plant siderophores. Siderophores from plants dising agent for silver [37], but recently it is being partly were described which also ¢t into the group of APCAs. replaced by L-ADA and PDTA [39]. An example is nicotianamine, an essential constituent of higher plants which is important for cellular iron transport 1.3.2. Use of APCAs to prevent catalysis mediated by and/or metabolism. Nicotianamine contains six donor metal ions groups (three nitrogen atoms and three carboxylate Small amounts of EDTA have been included in many groups) with an arrangement ideal for the formation of detergent formulations to stabilise the bleaching agent per- chelate rings [32]. Presently, there is no evidence that nic- borate by preventing metal-catalysed decomposition of the otianamine plays a role in iron uptake from the rooting compound [38]. In the bleaching process of the pulp and media, whereas this is known to be the case for the chemi- paper industry, hydrogen peroxide is increasingly em- cally similar mugineic acid and its acidic amino acid ana- ployed instead of chlorine compounds and the addition M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 77 of either EDTA or DTPA avoids the decomposition of 1.4.1. Concentrations of NTA and EDTA in sewage hydrogen peroxide catalysed by manganese or iron ions treatment plants [40]. APCAs are also used as additives for pharmaceuti- In Canada, where NTA accounts for up to 15 volume % cals, cosmetics and food to prevent transformation of the of laundry detergents, the typical NTA loading of raw ingredients or rancidity due to metal-catalysed reactions wastewater averaged 2500 Wgl31 [54]. Several case studies [37,41]. in Swiss municipal wastewater treatment plants revealed similar average concentrations in the in£uent of 100^1000 1.3.3. Use of APCAs to remove metal ions Wg NTA l31, whereas EDTA concentrations in raw waste- Multidentate chelating agents are widely applied in the water were determined to be somewhat lower ranging nuclear industry for decontamination of reactors and from 10 to 500 Wg EDTA l31 [55^57]. However, the equipment because they form water-soluble complexes EDTA concentrations in secondary e¥uents were approx- with many radionuclides. Although the exact composition imately ¢ve times higher than those of NTA, and did not of most of the decontamination reagents is proprietary di¡er signi¢cantly from the EDTA concentrations in the information, it is assumed that mainly NTA, EDTA, in£uent [57]. Obviously, a large portion of the incoming HEDTA or DTPA are present in these formulations NTA, but not of EDTA, was eliminated from the waste- [42]. In a single decontamination operation, hundreds of water during the treatment process. kg of chelating agents are employed [42,43]. The wastes obtained, containing radionuclide^chelator complexes, 1.4.2. Concentrations of NTA and EDTA in surface waters are solidi¢ed and disposed of [42^44]. Concern was raised Despite the dramatic increase of NTA usage in Canada that due to the presence of powerful chelators in such since 1970, no accumulation of NTA in Canadian surface wastes the migration of radioactive transition metals, waters was observed [58]. From 1971 until 1975, a median rare earths and transuranics may be enhanced [45]. concentration of 50 Wg NTA l31 was found in Canadian APCAs have also been proposed for application in the streams [59]. In Canadian coastal waters of the Atlantic remediation of metal-contaminated soils or sediments, ei- and the Paci¢c Ocean and in samples from the port area ther to serve as washing agents [46,47] or to support elec- of Halifax, no NTA was detectable more than 2 years trokinetic extraction processes [48,49]. Apparently, after its ¢rst usage in detergents [60]. APCAs have also the potential to be employed in phyto- The concentrations of NTA and EDTA found in Euro- remediation strategies due to their ability to increase metal pean rivers are in the range of 0^20 and 0^60 Wgl31, desorption from soil and to facilitate metal uptake by respectively [37,38,61]. For EDTA, however, occasionally plants [50]. Moreover, in medical treatment, CaEDTA is concentrations higher than 100 Wgl31 were found in Ger- used as antidote for the treatment of lead or other heavy man and Swiss rivers [62,63]. metal intoxications. EDTA complexes the toxic metal and In Swiss lakes, NTA concentrations measured were be- accelerates its excretion [51]. low 10 Wgl31. At the bottom of the lakes, NTA was present at approximately the detection limit of 0.1^ 1.3.4. Use of APCAs to increase metal availability 0.2 Wgl31. EDTA concentrations of 1^4 Wgl31 were found Since the early 1950s, synthetic chelating agents have throughout the whole waterbody and in contrast to NTA, been used to improve plant nutrition [52]. Especially the concentrations of EDTA were only slightly subject to APCA chelators are employed in fertilisers to supply temporal or spatial variations [61]. plants with trace metals such as iron, copper, zinc and In several sediment cores from Lake Greifensee, Swit- manganese. Most commonly, the Fe(III) chelates of zerland, EDTA concentrations in the range of 60 Wgkg31 EDTA, HEDTA, DTPA, EDDHA and ethylenediamine- to 1170 Wgkg31 were found [64]. The water content of the di(o-hydroxy-p-methylphenyl) acetic acid and the Cu(II), sediments was 80% with an EDTA concentration in the Zn(II) and Mn(II) chelates of EDTA are present in fertil- pore water equal to that found in the overlaying water isers [53]. column, namely around 6 Wgl31, indicating an enrichment of EDTA in the sediments of this lake. A similar enrich- 1.4. Concentrations of APCAs in the environment ment of EDTA in the sediment, ranging between 80 and 310 Wg EDTA kg31, was found in the Southern part of the Since the 1970s, NTA has received much attention be- Finish Lake Saimaa [65]. This lake receives a high load of cause of its use in laundry detergents and this has led to treated wastewaters from pulp and paper industries, which monitoring programs during which its concentration was use EDTA and DTPA in their production process [65, measured in various environmental compartments. More 66]. recently, interest was also focused on EDTA and the determination of EDTA concentrations was often included 1.4.3. Concentrations of NTA and EDTA in ground- and in the monitoring of NTA. In contrast, little is known drinking water about the occurrence and concentrations of other APCAs NTA concentrations determined in ground- and drink- in the environment. ing water ranged normally between 1 and 5 Wgl31 78 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

[58,61,67]. In Swiss groundwaters, EDTA concentrations 1.5. Speciation of APCAs in the environment of 0.1 to nearly 15 Wgl31 were found [61]. Drinking water obtained from surface water by bank in¢ltration from the In addition to environmental concentrations, informa- river Ruhr showed rather high EDTA concentrations (me- tion on the speciation of APCAs is essential to understand dian of 25 Wgl31), similar to those found in the supplying their environmental fate. So far, only data on the specia- river (median of 26 Wgl31). This indicates that no elimi- tion of NTA and EDTA have been reported. nation took place in the water along the in¢ltration pas- Considering typical concentrations of metal ions and sage. For NTA, however, the concentrations decreased by APCAs in natural systems, the chelating agents can be a factor of 10 during in¢ltration, resulting in a median expected to occur mostly as metal complexes. Unfortu- concentration of 0.7 Wgl31 in the drinking water [67]. nately, reliable analytical methods for monitoring the var- Due to the observed persistence of EDTA during in¢ltra- ious metal species of APCAs in natural waters are still tion, the compound was proposed to be used as tracer to lacking. Yet, in the case of EDTA, a method based on analyse how far in¢ltrated river water penetrates into high performance liquid chromatography has been devel- groundwater aquifers [68]. oped and ¢rst promising results have been obtained. How- ever, this method is still merely applicable to high concen- 1.4.4. Environmental concentrations of DTPA trations of the various EDTA species [71]. Attempts to Only little data are available on concentrations of exploit capillary electrophoresis have been made to assess DTPA in freshwater. In the few samples collected from the speciation of APCAs, in particular of EDTA and German rivers, DTPA concentrations between 2 and 15 DTPA. But this method is restricted to rather stable com- Wgl31 were determined [38]. Rather high concentrations plexes and its sensitivity is still too low for detecting com- for both DTPA and EDTA (10^20 Wgl31) were observed plexes in environmental samples [72,73]. Nonetheless, in the Southern part of the Finnish Lake Saimaa. In con- making use of the photolability of Fe(III)EDTA and the trast to EDTA, which was found in all samples, DTPA slow exchange kinetics of NiEDTA with Fe(III) ions, an- was only detected in the vicinity of the point sources and it alytical methods were developed to distinguish Fe(III)- was not found in the sediments despite its presence in the EDTA and NiEDTA from other EDTA complexes even overlaying waterbody [65,66]. Even though in these inves- when present at low, environmentally relevant concentra- tigations the analytical detection limit for DTPA was high- tions [7,64]. er than that for EDTA and the extractability from the Generally, equilibrium calculations based on the stabil- sediments might have been lower, this suggests that ity constants (see Eq. 1) have to be used to predict the DTPA was more readily degraded in this particular lake speciation of APCAs in ecosystems. However, two major water than EDTA. di¤culties a¡ect the prediction of APCA speciation under environmental conditions. Firstly, other natural ligands 1.4.5. Environmental concentrations of other APCAs including inorganic anions such as hydroxide, bicarbonate The occasional samples originating from various Ger- and phosphate anions and organic compounds such as man rivers revealed the presence of L-ADA and PDTA at humic acids compete with APCAs for the metals. Whereas concentrations around 1^5 Wgl31. In the same samples, concentrations of inorganic ligands can be measured and MGDA was not detectable. Furthermore, HEDTA was thus their e¡ect on APCA speciation can be predicted, it is measured at two sampling points in the river Elbe (Ger- more di¤cult to estimate the in£uence of organic ligands many) at rather high concentrations of about 40 Wgl31 on metal complexation. For marine systems, methods to [39,69]. These preliminary data indicate that it is not suf- analyse the complexation of a variety of metal ions with ¢cient to analyse only the `classical' representatives NTA, natural ligands were described [74^77]. For freshwater sys- EDTA and DTPA but that also other complexing com- tems, such methods exist for Cu2 and Zn2 [78,79] and pounds have to be considered when monitoring the pollu- Co2 and Cd2 [80,81]. The second di¤culty is that spe- tion of surface waters with APCAs. ciation calculations assume that the chemical equilibrium has been reached. However, and this applies in particular 1.4.6. Concentrations of APCAs in soils to EDTA, due to the slow kinetics of some of the metal We are not aware of reports on the concentration of exchange reactions, true speciation in a natural system APCAs in agriculturally used soils despite their usage in may di¡er considerably from the calculated equilibrium. fertilisers. Eventually, the highest environmental APCA Under conditions as they are found in natural waters, concentrations were reported for the Hanford site, USA, where a large excess of Ca2 and Mg2 over trace metals where wastes from nuclear decontamination processes exists, exchange reactions of EDTA complexes have been containing both complexing agents and radioactive metals shown to occur at slow rates with time scales of hours to were dumped into the ground [70]. In the mixed waste, days [82,83]. Especially Fe(III)EDTA was observed to ex- approximately 9 g EDTA l31, 10 g HEDTA l31 and 1^2 change rather slowly with other metals and for the ex- g NTA l31 were detected plus a wide variety of chelator change reactions with divalent cations, half-life times of fragments, also present in the mg^g l31 range. about 20 days were determined in river water [84]. Hence, M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 79

these values and the concentrations of natural ligands for Cu and Zn, the following distribution was predicted: 31% Fe(III)EDTA, 30% ZnEDTA, 15% Mn(II)EDTA, 12% CaEDTA, 10% NiEDTA, 2% PbEDTA and 0.5% CuEDTA [86]. Hence in contrast to NTA, co-ordination with heavy metals seems to be more important for EDTA, while the rather unstable CaEDTA complex makes up only a quite small portion of the total EDTA.

1.6. Environmental risks

The main concerns which have been raised in the dis- cussion on the risks of the occurrence of APCAs in the environment were: (1) adverse e¡ects on the operation of wastewater treatment plants, (2) toxic e¡ects of APCAs on aquatic and mammalian organisms, (3) the contribution of nitrogen from APCAs to eutrophication, and (4) the po-

Fig. 3. Speciation of EDTA and NTA in the Swiss river Glatt as pre- tential to mobilise metals. dicted by speciation calculations and, in case of Fe(III)EDTA and NiEDTA determined analytically. Data taken from [85]. 1. No negative e¡ects on the normal operation of wastewater treatment plants or sludge disposal systems have the complexation of EDTA in river water depends not been reported so far for both EDTA and NTA at en- only on the dissolved concentrations of the various cations vironmentally relevant concentrations [4,37]. and other ligands which determine the equilibrium speci- 2. Acute and chronic toxicity of NTA towards more than ation but also on the initial speciation of EDTA in which 50 species of freshwater and marine organisms has been it is released. This explains why Fe(III)EDTA makes up a studied [4]. Chronic `no observed e¡ect concentrations' considerable portion of total EDTA in surface waters, of NTA for aquatic life were at least one order of because it is mainly this species which is released into magnitude higher than measured environmental con- the environment from both photographic industry and centrations [4]. Furthermore, acute or chronic e¡ects wastewater treatment plants using iron salts for phosphate were only reported when the NTA concentrations precipitation. used in the tests were equal to or in excess of the con- Several authors presented speciation calculations for centration of divalent metal ions. In analogy, also NTA or EDTA, taking into account the above mentioned EDTA and DTPA were only weakly to moderately problems to a larger or smaller extent. Early predictions toxic for aquatic organisms when they were complexed for NTA speciation in freshwater including both inorganic with metal ions [37,87,88]. Given the fact that in sur- ligands and humic acids forcasted the chelating agent to face waters always a large stoichiometric excess of Ca be mostly bound to Cu2 when present at environmentally and other divalent metal ions is present and that the relevant concentrations of 10^20 Wgl31 [51]. According to actual NTA, EDTA and DTPA concentrations are more recent speciation modelling for a small river, which many orders of magnitude below the observed toxic receives an extensive load of treated wastewaters, NTA concentrations, no adverse e¡ects on aquatic life can present at 2 Wgl31 should be predominantly complexed be anticipated for these APCAs. with Ca2 (more than 90% of the total NTA) followed by In toxicity studies, NTA was only moderately toxic to Mg2 and Zn2 and only a small portion of NTA would mammals during acute oral exposure. It was not tera- be associated with other heavy metals (Fig. 3) [85]. The togenic itself or in presence of heavy metals and it was inclusion of natural ligands for Cu2 and Zn2 into these non-genotoxic. NTA was not metabolised but was rap- speciation predictions might explain the discrepancy with idly excreted by the kidney. Findings that urinary tract the earlier calculations. Apparently, the natural ligands for tumours can develop as a consequence of chronic ex- Cu2 compete successfully with NTA for the metal ion. posure to high doses of NTA were explained by The contradicting modelling data underline the need for changes in Zn and Ca distribution between urinary more information on the e¡ect of natural ligands on metal tract tissues and urine. But again, thresholds deter- speciation. mined for such e¡ects were much higher than human To determine the speciation pattern of EDTA in a river exposure resulting from the low environmental NTA system, a combination of measurements and equilibrium concentrations [4]. Also EDTA is only weakly toxic calculations was applied (Fig. 3). According to measure- to man (in fact, it is used as antidote and is permitted ments, the Fe(III)EDTA and NiEDTA fractions were set as food additive in some countries). WHO ¢xed the at 31% and 10%, respectively. Including into the model acceptable daily intake of EDTA to 2.5 mg kg31 80 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

body weight. Considering average concentrations of elimination has to be considered separately for each com- EDTA in drinking water, the daily uptake of EDTA pound. Numerous laboratory and ¢eld investigations is far below this value [37]. have shown that biodegradation is the key mechanism 3. It was suspected that widespread usage of NTA or for the removal of NTA from the environment [4], where- EDTA (both contain nitrogen) might enhance eutro- as EDTA is predominantly eliminated via photodegrada- phication. However, the contribution of both com- tion [62]. Before focusing on the biodegradation of pounds to the environmental nitrogen loading was APCAs, we will therefore summarise the most important found to be insigni¢cant for any direct eutrophication abiotic processes leading to either transformation of e¡ect [4]. It has also been discussed whether or not APCAs or to their elimination from environmental com- NTA or EDTA would stimulate algal growth indirectly partments. by extracting essential metals from sludges, sediments or humic acid and making them better available or ^ 2.1. NTA alternatively ^ by protecting organisms against the toxic e¡ects of certain heavy metals. While in labora- For both Fe(III)NTA and CuNTA, photochemical deg- tory systems indirect e¡ects were demonstrated at rela- radation by sunlight was reported, and this may ^ tively high NTA and EDTA concentrations, they are although only to a minor extent ^ contribute to the de- thought to be negligible in surface waters [4,37]. composition of NTA in the photic zones of lakes and 4. The potential of NTA and EDTA to alter heavy metal marine systems [4,93,94]. Photodegradation of other metal distribution has been thoroughly studied in model sys- complexes is not likely, since no signi¢cant decrease in tems (for reviews, see [4,37,89]). It was concluded that NTA concentrations was measured when solutions con- metal mobilisation by NTA and EDTA will not be taining NTA and an excess of Cd2,Pb2,Mg2 or signi¢cant at environmental concentrations and that Cr3 were exposed to light of 350 nm, a wavelength at variations in pH will have larger e¡ects on aqueous which both CuNTA and Fe(III)NTA were degraded. concentrations of heavy metal ions. However, shock The half-life for Fe(III)NTA during irradiation with sun- loadings in wastewater treatment plants with concen- light was determined to be approximately 1.5 h, whereas trations of free NTA or EDTA in the mg l31 range that for CuNTA was more than 100 times higher. The can cause mobilisation of Zn2 leading to elevated decomposition probably originates from a ligand to metal Zn2 concentrations in the e¥uent [55,90]. charge transfer resulting in a reduced metal ion and the Moreover, at pH 7^8 the mobilisation of toxic heavy formation of a ligand radical which then undergoes se- metals (e.g. Pb or Cd) sorbed to particles such as iron- quential decarboxylation giving rise to CO2 and formalde- (hydr)oxides resulted from an exchange of Fe(III) com- hyde, and iminodiacetate (IDA). Fe(III)IDA and Cu(II)- plexed by EDTA against the toxic metal, a reaction IDA were further photodegraded with glycine being which is catalysed by the particle surface [91]. Thus, formed. The latter, however, degraded more slowly than Fe(III)EDTA, which is an important EDTA species Fe(III)IDA, because of a shift of the absorption spectrum in many rivers, can remobilise adsorbed heavy metals of Cu(II)IDA towards shorter wavelengths and thus a during in¢ltration or in groundwater aquifers in calca- smaller overlap with the solar emission spectrum [95]. reous regions characterised by pH values s 7.0. In fact, such unwanted exchange reactions have already been 2.2. EDTA observed in a ¢eld investigation of an in¢ltration pas- sage [91]. On the other hand, in aquifers with lower pH The recalcitrance of EDTA towards biodegradation in values ( 6 7.0), Me(II)EDTA chelates will react with wastewater treatment plants or the environment has di- iron(hydr)oxides resulting in the dissolution of the min- rected much attention to other mechanisms of elimination. eral and the formation of Fe(III)EDTA. The metal lib- Thermic hydrolysis and indirect photolysis of EDTA are erated from the original EDTA complex will then ad- obviously negligible for its fate in natural systems [7]. Di- sorb to the surface and thus become immobilised. rect photodegradation, oxidation by metal(hydr)oxides Hence, under such conditions, no remobilisation but and, to a smaller degree, also sorption of EDTA to par- rather an immobilisation of toxic metals is to be antici- ticles and subsequent sedimentation of these EDTA- pated [92]. loaded particles seem to be important processes for the partial elimination of EDTA from aquatic systems.

2. Abiotic elimination of APCAs in the environment 2.2.1. Photolysis The process considered to be the most important for the The elimination of various APCAs from the environ- elimination of EDTA from surface waters is direct pho- ment is based on di¡erent biotic and abiotic processes. tolysis, which results from the fraction of sunlight below Despite the chemical and structural similarity of the 400 nm [62]. Apparently, only Fe(III)EDTA is susceptible APCAs, the mechanism primarily responsible for their to sunlight irradiation, whereas other environmentally rel- M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 81 evant EDTA species (complexes with Mg2,Ca2,Ni2, adsorbed EDTA fraction was quanti¢ed in some Swiss Cu2,Zn2,Cd2 and Hg2) will not photolyse. Under surface waters and wastewater treatment plants and it laboratory conditions, Mn(II)EDTA and Co(III)EDTA ranged around 1% of total EDTA, with 1036^1035 g also photodecomposed, although at rates approximately EDTA adsorbed to 1 g of particles [64]. Because of the one order of magnitude lower than that of Fe(III)EDTA negative charges of metal^EDTA complexes at pH values [96,97]. Interestingly, initially uncomplexed EDTA was typically found in natural waters (between pH 5.0 and pH photodegraded in the presence of lepidocrocite, an iron 8.0), the complexes sorb only to solids that are positively oxide [98], indicating that EDTA adsorbed to the surface charged, such as aluminium(hydr)oxides, iron(III)(hydr)- of the iron hydroxide can be photooxidised. Although free oxides and manganese(III,IV)oxides [7,103,104]. EDTA is not present in natural systems, the authors sug- Whereas adsorption onto such oxides does not play a gested that metal complexes might undergo photodegrada- role in the elimination of EDTA in wastewater treatment tion in an analogous manner by adsorbing to iron oxides plants and rivers, it could do so in lakes due to the sed- and forming ternary surface complexes. imentation of the particle-bound EDTA fraction [105]. Several researchers [40,99,100] have calculated the pho- This hypothesis is supported by the detection of enhanced tolysis half-lives of Fe(III)EDTA for surface waters at concentrations of EDTA in sediments compared to the various geographical locations. They ranged between water column and pore water ([64]; see also Fig. 7). only 11.3 min to more than 100 h depending on the light conditions employed for modelling. A comparison of ¢eld 2.2.3. Redox processes data with model calculations demonstrated that photolysis Lately, attention was drawn to the possible oxidation of by sunlight is really the most important process for the APCAs by +III metal oxidants in natural systems degradation of EDTA in the Swiss river Glatt [7]. While [106,107]. Such oxidants comprise manganese(III/IV)- on cloudy days no signi¢cant decrease in the EDTA con- (hydr)oxides, Co(III)-containing mineral phases (e.g. centrations could be detected along the river, all available Co(III) stemming from the oxidation of Co(II) adsorbed Fe(III)EDTA was eliminated by photodegradation under onto manganese(III/IV)- and iron(III)(hydr)oxides) as well sunny conditions within approximately 1 day. The remain- as iron(III)(hydr)oxides. While the potential of the Fe(III)/ ing EDTA found on sunny days consisted of photostable Fe(II) half reaction (E³ = +0.67 V) is too low to allow EDTA complexes. oxidation of EDTA, those for the Mn(III)/Mn(II) and As products of the photolysis of Fe(III)EDTA, ED3A, Co(III)/Co(II) half reactions (E³ = +1.50 V and E³= both possible isomers of ethylenediaminediacetate +1.48 V, respectively) are su¤ciently high. Indeed, inter- (EDDA) (N,N-EDDA and N,NP-EDDA), IDA and ethyl- actions of EDTA with Mn(III)OOH and Co(III)OOH enediaminemonoacetate (EDMA) were found, suggesting have been shown to result in the oxidation of EDTA, that the reaction mechanisms are similar to those already yielding the deacetylated breakdown products ED3A and described for the photolysis of NTA. Photolysis clearly EDDA. Mn(III) is a more reactive oxidant than Co(III) does not result in the total mineralisation of EDTA but and, therefore, the reactions of EDTA with Mn(III)OOH more easily biodegradable metabolites are formed during were signi¢cantly faster than those with Co(III)OOH. Ap- this process [101]. After a rather short irradiation (6.5 h) parently, EDTA does not persist in the presence of man- of a Fe(III)EDTA solution with sunlight, almost all of the ganese(III/IV)(hydr)oxides and reactions with such miner- initial EDTA was transformed into ED3A, EDDA and als were proposed to represent an important sink for EDMA. When this mixture of photolysis metabolites APCAs in the environment [107]. However, most experi- was incubated with activated sludge, a bioelimination of ments were done with free EDTA, and the in£uence of 53% was observed within 4 weeks. Using a Fe(III)EDTA metal complexation of the APCA molecule on the oxida- solution irradiated for 20 h which contained mainly tive process still has to be investigated in detail. EDDA and EDMA, bioelimination of the photodegradation products was 92%. These results indicate a higher 2.3. Other APCAs resistance of ED3A towards biological mineralisation when compared with EDDA and EDMA. Clearly, only Photodegradation was also reported for Fe(III)DTPA investigations with pure compounds can give a ¢nal an- but the products of the reaction were not determined swer. If ED3A actually turns out to be rather persistent, [40]. A theoretical half-life of Fe(III)DTPA lower than then further ¢eld investigations should be considered that of Fe(III)EDTA was calculated. Hence, the rates of which do not only include the monitoring of the disap- photodegradation of the iron(III) chelates apparently in- pearance of Fe(III)EDTA but also the fate of ED3A. crease in the order NTA 6 EDTA 6 DTPA. DTPA as well as HEDTA also adsorb to surfaces with 2.2.2. Sorption pH-dependent charge, i.e. aluminium- or iron oxides [108]. Several studies reported negligible adsorption of EDTA Yet, we are not aware of data on the adsorption of DTPA on humic acids, silica, kaolin, river sediments, humus sol- and HEDTA in aquatic systems and their role for the ids and activated sludge particles [55,56,102]. Recently, the elimination of these APCAs from the waterbody. 82 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

3. Microbial degradation of APCAs during wastewater population in coastal marine waters under aerobic or an- treatment and in the environment oxic conditions, several other researchers found NTA degradation in estuarine and o¡shore environments [122^124]. It has been demonstrated in numerous studies that NTA In particular, in samples from an estuarine system which is easily biodegradable during wastewater treatment, in had been pre-exposed to NTA for several years, NTA was natural waters and in soils. In addition, many investiga- rapidly degraded with no lag period at the low NTA con- tions indicate that biodegradation is complete and that centration of 10 Wgl31 [123]. IDA, a metabolite, which is formed intracellularly during Apparently, NTA degradation follows ¢rst order ki- the microbial breakdown of NTA, does not accumulate in netics, especially at the low NTA concentrations (below nature [4]. In contrast to NTA, contradictory results have 2 mg NTA l31) found in surface and groundwaters been published concerning the biodegradation of EDTA [125]. Similar rate constants (about 0.3 day31) were deter- or DTPA in the environment and only few reports can be mined for all freshwater systems situated in areas where found in the literature dealing with the biodegradability of NTA was not extensively used in detergents at the time of other APCAs. Hence, the question of the importance of the study. They were approximately 3^4 times lower than biodegradation for the environmental fate of these chelat- those found in a river from a region where NTA-containing agents is still unanswered. ing detergents were marketed. This again underlines the crucial role of acclimation of the microbial population 3.1. NTA present [123,126]. Similar to activated sludge from sewage treatment plants a few days to 4 weeks were typically 3.1.1. During wastewater treatment required for an unadapted microbial population to acquire During the treatment of wastewater, NTA is mainly NTA-degrading ability. Acclimation to NTA and subse- eliminated in biological treatment steps such as oxidation quent biodegradation occurred even at low environmental ponds and lagoons, activated sludge systems, or trickling concentrations between 5 and 50 Wg NTA l31 [126], but ¢lters. Systems that previously had not been exposed to the time period needed for developing degradation ability NTA frequently require an adaptation period of 1^4 seemed to decrease with increasing NTA concentrations, weeks before maximum degradation activity is achieved as tested for a concentration range from 0.02 mg NTA l31 [4]. Elimination e¤ciencies for NTA reported for di¡erent to 20 mg NTA l31 [127]. wastewater treatment plants range between 70 and s 90%. Analysis of measured NTA concentrations from a ¢eld Less e¤cient NTA removal (in some cases merely around study along a Swiss river receiving a high load of treated 50%) was only found at low temperatures during the win- wastewater indicated biological NTA elimination with a ter months or in treatment plants characterised by elevated half-life of about 8 h [62,128]. This agrees with the results sludge loading rates [54,109^115]. NTA elimination during of the laboratory studies described above [126]. Also the wastewater treatment was successfully simulated with a e¡ect of temperature on NTA degradation was investi- simple activated sludge model [116] which included only gated [62,126,129]. While in summer about 90% of NTA temperature-dependent growth of NTA-degrading micro- was eliminated from the river Glatt over a £ow distance of 31 organisms with NTA [WNTA(T) = 2.2 day exp(0.07 22 km, only 65% of the NTA disappeared in winter [62]. (T320³C))], but the model neglected decay of NTA de- Similarly, model calculations suggested that the behaviour graders and their growth with carbon sources other than of NTA in a Swiss lake can be attributed to biological NTA (the two e¡ects probably compensating for each degradation as only removal process with a constant deg- other). radation rate of 0.035 day31, corresponding to a half-life Due to lack of sorption, NTA is accumulated neither in of 20 days [105]. primary nor in secondary sludge and, therefore, the removal of NTA during anaerobic sewage sludge digestion 3.1.3. In aquifers and soils is not of utmost importance. Nevertheless, Klein [117] and In laboratory column systems with aquifer material Kirk and co-workers [118] demonstrated NTA reductions from a natural river water/groundwater in¢ltration site, in digesters varying from 8 to 45%, elimination of NTA NTA was rapidly mineralised under both aerobic and de- close to 100% was observed when activated sludge already nitrifying conditions [130]. According to these data, it was acclimated to NTA was added to the digestion process expected that very low NTA concentrations ( 6 2 Wgl31) [119]. will be found in the groundwater even in cases where the in¢ltration distance was only a few meters. Hence, even a 3.1.2. In the aquatic environment signi¢cant increase in NTA concentration in river waters According to numerous laboratory studies, NTA is bio- should not lead to higher concentrations in the ground- logically degraded in freshwater [4,120]. Contradictory re- water. Low NTA concentrations measured in various sults have been reported on its degradation in marine and groundwater samples, also in those obtained from in¢ltra- estuarine water samples. Whereas Kirk and co-workers tion areas in close distance to the feeding river, con¢rmed [121] found no NTA degradation by a marine bacterial these predictions [38,61,67,131]. M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 83

Also in soils, NTA readily decomposed under aerobic Later, Gschwind [144], working under aerobic condi- conditions with half-lives ranging from 3 to 7 days [132^ tions with micro-organisms ¢xed on carrier materials, ob- 134]. In contrast to Tiedje and Mason [132], both Taba- served degradation of EDTA leading to the formation of tabai and Bremner [135] and Ward [133] observed NTA nitrate (probably because of the presence of nitri¢ers). degradation also under anaerobic conditions where nitrate With this system, treatment of a model wastewater con- substituted molecular oxygen as a terminal electron accep- taining 200 mg EDTA l31 under continuous conditions led tor. While nitrate concentrations in the system had no to 99% elimination at a hydraulic retention time of 1.5 h. in£uence on the initial rates of NTA degradation, they Addition of yeast extract or glucose did not a¡ect the a¡ected the extent of the mineralisation under anaerobic performance. Interestingly, the optimum pH for e¡ective conditions [133]. EDTA degradation was rather high (between pH 9.0 and 9.5). A similarly high optimum pH for EDTA degradation 3.2. EDTA was reported by Van Ginkel and co-workers [145] in semi- continuous activated sludge units which were operated at 3.2.1. During wastewater treatment and in laboratory test pH 8^9, whereas at pH 7.0, no EDTA removal was de- systems tected even after a long incubation period of 6 months. Balances of the EDTA load in the in£uent and e¥uents Hence, EDTA degradation seems to be favoured at mod- of municipal wastewater treatment plants with sampling erately alkaline conditions. Also, in sewage treatment periods of several days gave no indication for a signi¢cant plants for wastewaters from pulp and paper industries, elimination of EDTA neither by biological nor by physi- an increase of EDTA elimination e¤ciencies from 10 to cochemical processes [55,56,136,137]. At the same time, 50% was observed when the operating pH was raised from NTA was e¤ciently eliminated ( s 85%) in all plants in- the range between 6.5 and 7.0 up to 8.0 or even 9.0 vestigated. Also, several laboratory studies with inocula [146,147]. Increasing the pH at which a sewage treatment collected from industrial or municipal wastewater treat- unit is operated to values between 8.5 and 9.0 to achieve ment systems consistently demonstrated the recalcitrance the degradation of EDTA in activated sludge in the ab- of EDTA [138^141]. However, e¤cient EDTA elimination sence of support material has now even been patented of about 80% in an industrial wastewater treatment plant [148]. Under these conditions, also PDTA was degraded, has been reported lately [142]. In activated sludge samples yet, rather long sludge retention times of at least 2 months taken from the very plant, almost 100% of added EDTA were needed for e¤cient PDTA breakdown, while for (measured as DOC) was degraded within 10 days. EDTA removal about 1 week was su¤cient. Several reports have already described biologically Eventually, Thomas and co-workers [149] isolated a mediated EDTA degradation under laboratory conditions. mixed, EDTA-degrading culture from river water and The ¢rst report [143] demonstrated decomposition of sludge from an industrial wastewater treatment plant. EDTA by microbial populations from an aerated lagoon Among the cultivatable micro-organisms within this cul- receiving EDTA-containing industrial e¥uents. The au- ture, representatives of the genera Methylobacterium 14 thors followed [ C]CO2 formation from the iron(III) com- (35%), Variovorax (17%), Enterobacter (15%), Aureobacte- plex of radioactively labelled EDTA when incubated in the rium (11%) and Bacillus (11%) were predominant. How- dark to prevent photodegradation. After an incubation ever, the ability of the isolated bacteria to grow with period of 5 days, about 90% of the initially present EDTA was not tested. Growth of the culture with EDTA had disappeared. 27% of the initial radioactivity Fe(III)EDTA was slow, with a doubling time of 66 h, of the acetate-labelled and 31% of the ethylene-labelled and only 60% of the initial Fe(III)EDTA (5 mM) was 14 EDTA was recovered as CO2, indicating that both the degraded. The mixed culture grew also with possible met- ethylene backbone and the acetyl groups were attacked. abolic intermediates such as EDDA, ethylenediamine, During EDTA degradation, the concentrations of ED3A NTA and IDA. From this it was concluded that build- and IDA increased transiently. Most likely, these two up of these compounds was unlikely during growth with compounds were intermediates of the microbial EDTA EDTA. degradation. In contrast, other possible intermediates There is only one study analysing the behaviour of such as NTA, N,N-EDDA and N,NP-EDDA, EDMA EDTA in a marine ecosystem. In microcosms containing and glycine were found only at very low concentrations. sea water and sediment illuminated with UV light, about Supplementing the culture with NTA or ethylenediamine 50% of the initially added Fe(III)EDTA complex was con- resulted in a stimulation of EDTA degradation whereas verted after 17 weeks [150]. Unfortunately, the poor ex- the addition of amino acids and sugars had an inhibitory perimental set-up used does not allow clear distinction e¡ect. Unfortunately, the study was not extended to the between biodegradation and photodecomposition of the fate of EDTA in the lagoon from where the culture was Fe(III)EDTA chelate. obtained to give an indication whether or not the consortium was also consuming EDTA under environmental 3.2.2. In soils and sediments conditions. Contradictory results have been published concerning 84 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

EDTA degradation in soils and sediments. Some groups sured HEDTA concentration was below 5% of that ini- found no biological EDTA breakdown [151,152], whereas tially supplied. In sterile controls, the fraction of non- others observed a slow microbial EDTA decomposition biologically degraded HEDTA in this experiment was de- under aerobic conditions [153^156]. No EDTA mineralisa- termined as 27%. Nevertheless, recently HEDTA was con- tion was found under anaerobic conditions [154,156]. sidered to be biologically not degradable [160]. Typical biodegradation tests using activated sludge as 3.3. DTPA an inoculum and aerobic conditions showed that the amino acid derivatives L-ADA, SDA and MGDA are readily 3.3.1. Wastewater mineralised by micro-organisms. Biodegradation of ASDA From measurements of EDTA and DTPA concentra- was found to be stereospeci¢c and only L-ASDA was tions in wastewater e¥uents from pulp and paper mills, easily biodegradable [9]. in the in£uent and e¥uent of wastewater treatment plants of these mills [157] as well as in the receiving lake [66], it 3.5. Naturally occurring APCAs was inferred that DTPA is biologically and/or chemically more readily degradable than EDTA. Indeed, in an acti- 3.5.1. Biodegradation of EDDS vated sludge process, a reduction of DTPA amounting to In some applications, EDDS is suggested to replace the 50^70% was observed while only 30% of EDTA was elim- poorly degradable EDTA. Several biodegradation studies inated from the wastewater. Also in model wastewater with activated sludge from di¡erent sources have consis- treatment plants, the initial DTPA concentration was re- tently demonstrated the stereospeci¢city of EDDS break- ported to be reduced by about 50% [158]. However, elim- down [141,161]: whereas [S,S]-EDDS and [R,S]-EDDS ination could not unequivocally be attributed to microbial disappeared rapidly, [R,R]-EDDS was recalcitrant. How- activity. In contrast, Hinck and co-workers [140] did not ever, only the S,S-isomer was completely mineralised [161] observe DTPA breakdown in activated sludge although and the transformation of [R,S]-EDDS led to the produc- the inoculum for the biodegradability tests was taken tion of a recalcitrant intermediate (AEAA), indicating the from a treatment plant that had been in contact with removal of one succinyl residue only. This suggests that DTPA for over 5 years. only [S,S]-EDDS should be employed for the application in both domestic and industrial products. 3.3.2. Soils and sediments Takahashi and co-workers [141] investigated the degra- In soils and sediments, several groups observed the mi- dation of propanediaminedisuccinate (PDDS), a com- crobial breakdown of DTPA [151,155,156,159] and only pound very similar to EDDS and also occurring as three Allard and co-workers [152] were unable to detect disap- di¡erent stereoisomers. In contrast to EDDS, all three pearance of DTPA in sediments. Both NTA and EDTA isomers of PDDS were easily biodegradable. However, it were formed as products of DTPA breakdown apparently was not determined whether the compound was com- resulting from the cleavage of a C^N bond within one of pletely mineralised. the ethylenediamine parts of the DTPA molecule [151]. Further metabolites were identi¢ed as incompletely substi- 3.5.2. Other naturally occurring metal-chelating compounds tuted APCAs such as diethylenetriaminetetraacetate, di- There is little information on the possible biodegrada- ethylenediaminetriacetate and ED3A which will spontane- tion of naturally occurring APCAs. Nevertheless, several ously cyclise under neutral and especially acidic conditions studies indicate that plant-produced chelating organic to form oxopiperazinepolycarboxylic acids [159]. Unfortu- compounds, such as the APCA mugineic acid, are de- nately, it was not possible to distinguish whether these graded by rhizosphere micro-organisms under environ- cyclised substances were formed during DTPA transfor- mental conditions [162^165]. In this way, the microbial mation itself or only subsequently during the analytical £ora may reduce iron uptake of plants by both the deg- procedure. Nevertheless, the authors proposed that they radation of the phytosiderophores and by competition for might be formed during DTPA degradation and then ac- iron. cumulate in the environment due to their high stability. Some of these metabolites were also detected in samples taken from di¡erent German rivers during a subsequent 4. Degradation of APCAs by pure microbial cultures screening program. The dominant metabolite was ED3A and its corresponding cyclised form [159]. Up to now, successful isolation of pure, APCA-degrading microbial cultures has been restricted to microbial 3.4. Other synthetic APCAs strains that are able to grow with NTA or EDTA. Some of these isolates have lately been shown to also be able to Only one report is available which describes the slow utilise [S,S]-EDDS (see Section 5.3). Reports on strains degradation of HEDTA by soil micro-organisms [155] ac- utilising DTPA or HEDTA, two other industrially impor- cording to which after 173 days of incubation the mea- tant APCAs, have not been published so far, and we are M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 85 not aware of any attempts to isolate micro-organisms that three distinct groups (Table 3): the ¢rst group contains grow with representatives of the new generation of APCAs obligately aerobic, motile rods which are mostly pleo- such as L-ADA or MGDA. morphic and can utilise sugars and also a wide range of other substrates including methylated amines, hence 4.1. Pure cultures utilising NTA indicating their ability to assimilate C1 units. These strains are phylogenetically localised in the Agrobacterium^Rhi- The isolation of NTA-degrading micro-organisms from zobium branch of the K-subclass of Proteobacteria and a wide range of ecosystems has been described including they have been combined to form the new genospecies river waters, activated sludge, sediments and soils [166^ Chelatobacter heintzii [181]. Most recent data indicate 178]. Most of the pure cultures consisted of Gram-nega- that Chelatobacter strains are closely related to bacteria tive, obligately aerobic rods capable of utilising NTA as from the genus Aminobacter which are able to grow with sole source of carbon/energy and nitrogen. Although for methylated amines [183,184]. It is therefore possible that in the majority of Gram-negative strains detailed character- the future both groups will be combined to a single genus isation was missing, reports before 1985 mostly allocated [185]. them to the genus Pseudomonas [166^170,172^175]. One The second group of isolates (Table 3) consisted of ob- Gram-negative strain was identi¢ed as Listeria sp. [139]. ligately aerobic, non-motile short rods or diplococci un- Two Gram-positive isolates were described and one of able to grow with sugars. They again could not be as- them was identi¢ed as a Bacillus sp. [175], whereas for signed to an existing genus and, therefore, were strain TE3 [177] further work is needed to de¢ne its exact established as the new genospecies Chelatococcus asaccha- taxonomical position (morphological and physiological rovorans [181]. This genus belongs to the K-subclass of studies indicate its a¤liation to the Corynebacterium sensu Proteobacteria and 16S rRNA sequence comparison indi- stricto, Mycobacterium, Nocardia, Rhodococcus group cated Rhodopseudomonas as nearest neighbour. [179]). In addition to the strictly aerobic bacteria, denitri- The range of APCAs, which can be used as growth fying strains have also been isolated [171,178], whereas ^ substrates by Chelatococcus and Chelatobacter strains, so far ^ the enrichment of sulfate-reducing NTA degraders comprised NTA and IDA, while EDTA did not support has not been successful [171,180]. growth [177]. Moreover, C. asaccharovorans grew with A more detailed taxonomical study of Gram-negative [S,S]-EDDS, whereas Chelatobacter strains did not [186]. NTA-utilising isolates [177^179,181], including most of No further APCAs have so far been tested as growth the isolates still available at the time, revealed that none substrates for the NTA-degrading strains. of them can be assigned to the genus Pseudomonas as The third group of NTA degraders presently consists of de¢ned by de Vos and de Ley [182]. These isolates formed only one strain (TE11) that is facultatively denitrifying

Table 3 Some key properties of the best investigated pure bacterial cultures able to utilise APCAs as a sole source of carbon, energy and nitrogen

31 Isolation Strain [reference] Taxonomical position Main properties Respiration Wmax (h )on Growth with substrate isolation APCAs other substrate than isolation substrate NTA C. asaccharovorans strain K-2 branch of Gram-negative, diplococci, obligately 0.08^0.1 [S,S]-EDDS TE1 and TE2 [181] Proteobacteria non-motile, with S-layer, aerobic not utilising sugars C. heintzii, strains K-2 branch of Gram-negative, rods, obligately 0.1^0.15 IDA TE4^TE10 and ATCC Proteobacteria, within the pleomorphic, motile (1^3 aerobic 29600 and ATCC Agrobacterium^Rhizobium subpolar £agella) 27109 [181] branch strain TE11 [178] Q-branch of Proteobacteria, Gram-negative, rods, aerobic and 0.08 (aerobic); closely related to motile denitrifying 0.03 Xanthomonas (denitrifying) EDTA A. radiobacter ATCC Gram-negative, rods aerobic PDTA (only 55002 [190] degrading EDTA only when complexed when complexed with with Fe(III)) Fe(III) strain BNC1/DSM K-branch of Proteobacteria Gram-negative, rods, aerobic 0.024 NTA, IDA, ED3A 6780 [191] non-motile when grown with EDTA strain DSM 9103 [192] K-branch of Proteobacteria Gram-negative, diplo-rods, aerobic 0.06^0.07 NTA, IDA, within the Agrobacterium^ non-motile N,NP-EDDA, Rhizobium branch [S,S]-EDDS 86 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

(Table 3). Apart from its ability to denitrify and its inabil- for several years. Consequently, the isolate grew with ity to grow with IDA, this Gram-negative motile isolate Fe(III)EDTA as sole source of carbon, whereas no signi¢- di¡ers only slightly from C. heintzii with respect to its cant growth was observed with uncomplexed EDTA [190]. physiological and nutritional characteristics. Polyamine The strain was able to degrade Fe(III)EDTA at initial and quinone patterns, however, discriminate TE11 from EDTA concentrations higher than 100 mM but it failed the latter and suggest an allocation to the Q-subclass of at low concentrations, and the residual EDTA concentra- Proteobacteria. The presently available data suggest that tions were always in the range of approximately 3^5 mM. TE11 is closely related to but still clearly distinct from In batch culture, the extent of degradation was more com- members of the genus Xanthomonas [178]. plete for lower initial pHs (probably due to a slower in- Surface antibodies were raised against C. heintzii and crease of culture pH from excreted NH3 during EDTA C. asaccharovorans and were used to investigate their dis- breakdown). Addition of peptone or yeast extract to the tribution in the environment [187,188]. In surface waters, medium did not a¡ect EDTA degradation but that of cells reacting with antibodies made up for 0.01^0.1% of glycerol reduced the amount of EDTA degraded drasti- the total (acridine orange) cell number. In activated cally. It was suggested that in the presence of glycerol, sludges of wastewater treatment plants, the proportion the Agrobacterium strain utilised EDTA merely as nitro- was about 10-fold higher. Surprisingly, no signi¢cant dif- gen source and no longer as carbon and energy source. ference in the fraction of cells reacting with antibodies was Other APCAs were tested for growth of the Agrobacterium observed between samples collected from treatment plants, strain, but only Fe(III)PDTA was degraded and no which di¡ered considerably with respect to their e¤ciency growth was observed with NTA, IDA or EDDA, although in NTA elimination [187]. Despite the immunological de- the latter compounds are supposed to be intermediates in tection of bacteria of the genera Chelatobacter and Chela- the breakdown of EDTA [143,193]. Furthermore, the tococcus in activated sludge and in the environment, the strain was able to utilise a wide range of sugars, carboxylic contribution of these strains to the degradation of NTA is acids and amino acids as growth substrates. still unknown. To tackle this question, an experimental From an industrial sewage treatment plant receiving approach would be required which consists of the testing EDTA-containing wastewater, No«rtemann [191] was able of individual cells present in environmental samples for to isolate a mixed culture of strain BNC1 and BNC2 that both their serological reaction with the antibodies above grew at considerably lower initial EDTA concentrations described and for NTA degradation. (91 mM) than the Agrobacterium strain. In the mixed It should be added here that in several Pseudomonas culture, only strain BNC1 degraded EDTA and its meta- strains (P. aeruginosa, P. £uorescens and P. cepacia), bolic activity was stimulated in the presence of other mi- NTA stimulated growth by promotion of the iron uptake, cro-organisms probably providing vitamins. Indeed, addi- although it did not serve as growth substrate [189]. Ap- tion of vitamins was demonstrated to signi¢cantly enhance parently, NTA assisted the intracellular incorporation of growth of a pure culture of strain BNC1 [194]. As long as iron through an active transport system, as do naturally EDTA was complexed with alkaline earth metal ions, occurring siderophores. In contrast, other chelators such strain BNC1 tolerated EDTA concentrations up to ap- as EDTA, EDDS and EDDHA had an inhibitory rather proximately 16 mM. However, the presence of uncom- than a stimulating e¡ect on growth of the tested bacterial plexed EDTA considerably reduced the viability of cells strains. Hence, in the environment, NTA might act as a and no growth with free EDTA was observed [194]. A siderophore-like compound for some micro-organisms. similar observation had already been reported for an However, the interesting question of the intracellular fate NTA-degrading strain [176]. Such adverse e¡ects are prob- of NTA used by micro-organisms as a siderophore has not ably due to interactions of free EDTA or NTA with metal been further studied. ions that stabilise the bacterial cell surface because destabilising e¡ects of free EDTA on various Gram-negative 4.2. Pure cultures utilising EDTA bacteria are well known [195,196]. A taxonomic analysis indicated an allocation of strain BNC1 to the K-subgroup Most likely because of the poor biodegradability of of Proteobacteria [193]. However, a more detailed taxo- EDTA, the successful isolation of EDTA-degrading bac- nomical study including 16S rRNA sequencing is needed teria has only recently been reported [190^192] (Table 3). to more exactly allocate this strain, in particular to deter- While the isolates described by No«rtemann [191] and Wit- mine its relation to the other EDTA-degrading strains schel and co-workers [192] exhibit many similarities, the known so far. pure culture isolated by Lau¡ and co-workers [190] has Recently, another EDTA-degrading bacterial isolate rather di¡erent properties with respect the EDTA degra- DSM 9103 has been described [192]. Starting with the dation characteristics. mixed culture from Gschwind [144] a Gram-negative, ob- The latter strain, identi¢ed as Agrobacterium radio- ligately aerobic bacterium was obtained which also be- bacter, was isolated from a secondary waste treatment longs to the K-subgroup of Proteobacteria. Both polar facility that had received wastes containing Fe(III)EDTA lipid pattern and 16S rRNA sequence of the strain are M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 87 indicative of members of the Agrobacterium^Rhizobium ably as a result of changes in the composition of the mi- branch. Indeed, the 16S rRNA sequence of the isolate crobial population. In fact, the bio¢lm on the carrier par- showed highest similarity (96.1%) to that of Rhizobium ticles was found to become more ¢lamentous and loti [186], now renamed Mesorhizobium loti [197]. When voluminous when feeding the substrate mixture. Hence, added to raw wastewater or activated sludge from a mu- in wastewaters containing a high fraction of readily de- nicipal wastewater treatment plant cells strain DSM 9103 gradable carbon, a two-step process might be of advantage were able to degrade EDTA to completion (initial concen- where ¢rst the easily degradable carbon sources are elim- tration approximately 3.5 mM) [192]. However, supple- inated before EDTA is removed with the help of strain mentation with mineral medium (containing high concen- BNC1 in a second step [212]. trations of alkaline earth metal ions) was necessary to Industrial wastewaters often contain the Fe(III) com- obtain e¤cient EDTA removal in activated sludge, indi- plex or other heavy metal complexes of EDTA, but strain cating again the importance of EDTA complexation for its BNC1 is not able to attack these chelates (see Section 7). biodegradability. Interestingly, strain DSM 9103 is not To treat wastewaters containing such EDTA complexes, a only able to grow with EDTA but also with other APCAs precipitation step was established before the EDTA solu- such as NTA, IDA and [S,S]-EDDS [198]. tion was supplied to the bioreactor. By the addition of a Although ferrichrome and ferrioxamine type sidero- large excess of calcium hydroxide to a Fe(III)EDTA-con- phores do not belong to the class of APCAs, it should taining synthetic wastewater, Fe3 was exchanged against be mentioned here that pure cultures of ferrichrome A- Ca2 and then precipitated as iron hydroxide. A solution degrading micro-organisms had been isolated already in containing mainly CaEDTA was obtained, and after ad- the 1960s and that the breakdown of these siderophores justing the pH, the Fe(III)EDTA elimination e¤ciency by has been studied [199^207]. Interestingly, based on 16S strain BNC1 was about 98% [213]. rRNA homology, a deferrioxamine B-degrading bacterium Eventually, the system was tested for its performance in was found to be a strain of R. loti [205] and, therefore, it eliminating EDTA in real wastewaters originating from seems a close relative of some of the EDTA-degrading the dairy industry, vat cleaning in a brewery, the pulp strains presently known (see above). Speciation was found and paper industry, and from the photographic industry. to be a¡ecting degradability with only the iron-free com- The data indicated that all these wastewaters were treat- pound (deferrioxamine B, but not ferrioxamine B) being able with immobilised EDTA-degrading bacteria but occa- degraded [202,203,206]. sionally speci¢c pre-treatment steps were required. EDTA in wastewater from pulp and paper manufacturing was 4.3. Potential of bacterial isolates for the treatment of e¤ciently removed without any pre-treatment. Waste- APCA-containing wastewaters waters from the dairy industry could be treated in a two-step process because of the presence of high amounts NTA is removed e¤ciently during biological wastewater of non-EDTA DOC, which must be removed ¢rst. Waste- treatment (see Section 3) and, therefore, the application of waters from vat cleaning or the photographic industry NTA-degrading isolates for a special treatment of NTA- again contain high amounts of heavy metals and a precip- containing wastewaters has not been investigated. In con- itation step (or a similar process to remove these metals) trast, elimination of EDTA in most wastewater treatment has to be carried out before biological treatment is possi- plants is negligible [55,56,137]. Thus, the use of EDTA- ble [212]. The results are quite promising with respect to a degrading strains for the treatment of wastewaters con- potential utilisation of immobilised cells of strain BNC1 taining high concentrations of EDTA was proposed and for the biological treatment of EDTA-containing waste- also patented [208^212]. waters. Nonetheless, a note of caution should be added A process based on a mixed culture of strain BNC1 and because all tests for EDTA degradation in industrial BNC2 for treating industrial EDTA-containing waste- wastewaters have so far been performed under constant waters has been developed. The cells were ¢xed on carrier conditions, a situation that is rarely encountered in real material and employed in a bio¢lm airlift-loop reactor life. Therefore, important data are still missing with re- [211]. The reactor was operated at dilution rates between spect to overall elimination e¤ciencies in case of varia- 0.06 and 0.3 h31 and allowed to achieve an EDTA elim- tions in wastewater composition or other £uctuations of ination e¤ciency between 95 and 98% for an EDTA con- parameters (e.g. in pH or feeding rate). centration of 0.45 g l31 in the feed medium. To mimic the situation in real wastewater treatment plants where EDTA is rarely the sole source of carbon and nitrogen, a mixture 5. Biochemistry and genetics of APCA degradation of EDTA and glycerol (each 300 mg l31) was fed to the reactor (D = 0.2 h31). Initially, when feeding the mixture, Information on the biochemistry and genetics of APCA no signi¢cant reduction of EDTA elimination e¤ciency degradation is of course restricted to those compounds for was observed but after more than 10 days, the extent of which isolation and cultivation of pure bacterial cultures EDTA degradation started to decrease gradually, prob- has been reported, i.e. NTA, EDTA and [S,S]-EDDS. 88 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Again, of all APCAs, NTA has clearly received most at- NTA degradation and of enzymatic activities enhanced in tention. cell-free extracts of NTA-grown cells, already some 25 years ago a pathway for the aerobic NTA breakdown 5.1. NTA degradation was proposed (Fig. 4) [166,170,174,215]: NTA was thought to be metabolised in two steps via IDA to glycine Research mainly focused on the intracellular metabo- and glyoxylate, with the ¢rst step being catalysed by a lism of NTA and the question of its transport into cells monooxygenase (MO) [166,170]. The products were then was rarely addressed. One can merely speculate that NTA channeled into the central metabolism. A comparison of is actively transported across the cell membrane because enzyme activities in cell-free extracts of glucose- and NTA- [14C]NTA uptake was inhibited by KCN [169,214]. grown cells gave strong evidence for the involvement of glycine decarboxylase and serine hydroxymethyl transfer- 5.1.1. NTA catabolism ase in the subsequent metabolism of glycine, and that gly- Based on identi¢cation of products accumulated during oxylate carboligase and tartronic semialdehyde dehydroge-

Fig. 4. Metabolic pathway of NTA in obligately aerobic and facultatively denitrifying Gram-negative bacteria. (1) Transport enzyme; (2) NTA MO; (3) NTA DH; (4) NTA DH/NtrR complex; (5) IDA DH; (6) glyoxylate carboligase; (7) tartronate semialdehyde reductase; (8) serine hydroxymethyl transferase; (9) glycine synthetase (decarboxylase); (10) serine:oxaloacetate aminotransferase; (11) transaminase; (12) glutamate DH; (13) hydroxypyru- 5 10 vate reductase; (14) glycerate kinase; (methylene)FH4,(N,N -methylene) tetrahydrofolic acid; PLP, pyridoxal phosphate; TPP, thiamine pyrophosphate (from [6] with permission). M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 89

nase (DH) took care of the transformation of glyoxylate from NADH2 to FMN (Fig. 5). cB can be replaced by into glycerate [170]. other NADH2 :FMN oxidoreductases, for instance the The proposed pathway was con¢rmed by puri¢cation oxidoreductase from Photobacterium ¢scheri which nor- and characterisation of two enzymes from C. heintzii mally supplies FMNH2 to the luciferase enzyme. There- ATCC 29600 that catalyse the degradation of NTA and fore, FMN and FMNH2 most presumably act here as real IDA, respectively [216,217]. A MO (NTA MO) was found coenzymes not being tightly bound to cB. Recently, sim- to catalyse the attack on NTA leading to the formation of ilar two-enzyme systems have been described, also consist- IDA and glyoxylate. In a second step, IDA is oxidatively ing of a NADH2 :FMN oxidoreductase and a MO, namely split into glycine and glyoxylate, not by NTA MO as those involved in the oxidation of dibenzothiophene [221] originally supposed [166], but by a membrane-bound and in the synthesis of antibiotics such as pristinamycin IDA DH. Both enzymatic activities were not only detected IIA [222], actinorhodin [223] or valanimycin [224]. It there- in C. heintzii but also in Chelatococcus strains [218]. fore is not surprising that cA and cB exhibited the highest homology with proteins of this two-enzyme system group 5.1.1.1. Characteristics of NTA MO. NTA MO from when the amino acid sequences were compared [220, C. heintzii ATCC 29600 was ¢rst puri¢ed and character- 224,225]. Eventually, sequence comparison revealed a sig- ised by Uetz and co-workers [216] and later more infor- ni¢cant similarity between FMNH2-dependent MOs from mation was obtained by sequencing the corresponding E. coli, Bacillus subtilis and Pseudomonas putida catalysing genes [219,220] (Table 4). NTA MO was originally the desulfonation of alkanesulfonates and several of the thought to consist of two weakly associated components, above mentioned MOs (including NTA MO) [226]. In this cA and cB, both of which must be present to catalyse the study, also a conserved domain present in all MOs was transformation of NTA [216]. More recently, it was dem- detected which might be part of the active site of these onstrated [220] that cA and cB are two distinct enzymes, enzymes. This domain, however, contained none of the with cA being the true MO catalysing the oxidative cleav- protein motifs known so far. age of NTA with the consumption of FMNH2 and mo- Also from Chelatococcus strain TE2, a FMNH2-depen- lecular oxygen, and cB being an oxidoreductase providing dent MO was puri¢ed which exhibited many similarities FMNH2 for the MO by transferring reducing equivalents with cA from C. heintzii ATCC 29600. The N-termini of

Table 4 Some properties of key enzymes involved in APCA degradation NTA MO (cA) from NTA DH from IDA DH from EDTA MO (cAP) EDDS lyase C. heintzii strain TE11 C. heintzii from strain from strain DSM 9103 DSM 9103 Location Soluble Soluble Membrane-bound Soluble Soluble Molecular mass SDS (kDa) 47 80 ^ 44 58 Native (kDa) 99 170 ^ 210 130 Substrate NTA NTA IDA EDTA [S,S]-EDDS, [R,S]-EDDS Products IDA+glyoxylate IDA+glyoxylate Glycine+glyoxylate ED3A, N,NP- AEAA+fumarate EDDA+glyoxylate Further substrates ^ ^ ^ DTPA, HEDTA, ^ NTA, PDTA a Cofactors required FMNH2 PMS ^ FMNH2 ^ Redox component ^ FAD (cyt b)b ^^

Electron acceptor O2 PMS, in vivo acceptor Respiration chain, O2 ^ unknown ubiquinone pool

Accessory enzymes NADH2 :FMN NtrR (under anoxic NADH2 :FMN oxidoreductase (cB) conditions) oxidoreductase (cBP) Required speciation MgNTA (CoNTA Uncomplexed NTA Uncomplexed IDA MgEDTA (ZnEDTA, Uncomplexed EDDS of substrate ZnNTA, MnNTA, (enzymatic degradation MnEDTA, CoEDTA) (MgEDDS, BaEDDS, NiNTA, Fe(III)NTA) also in presence of CaEDDS, MnEDDS) Mg, Ca, Ba and Mn)

Km(substrate) 0.5 mM (for MgNTA) 0.095 mM (for NTA 8 mM (in presence 0.8 mM (for MgEDTA) 0.2 mM DH/NtrR complex); of Ca) 0.19 mM (for NTA DH only, under aerobic conditions) Data were taken from [198,216,217,220,229,235,251,252]. aIn vivo cofactor unknown. bIndicated by spectral properties. 90 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 both proteins were virtually identical with only one amino IDA DH [217] (Table 4). IDA DH from C. heintzii was acid being di¡erent. In addition, the overall amino acid distinctly di¡erent from other membrane-bound DHs such composition of the two MOs was very similar (however, as succinate DH and it merely catalysed the oxidative the whole amino acid sequence of cA from C. asaccharo- splitting of IDA to glycine and glyoxylate. IDA DH is vorans has not been unraveled yet). In contrast, no protein probably not only associated with the membrane but ex- equivalent to cB has yet been isolated from C. asaccha- ists as an integral membrane protein which feeds electrons rovorans, but the existence of such a protein is compulsory from the oxidation of IDA into the electron transport as indicated by the fact that cB from C. heintzii could chain via the ubiquinone pool [217]. Partial enrichment form a functional, NTA-oxidising enzyme complex with of IDA DH was achieved by extracting the enzyme from cA from C. asaccharovorans [218]. membranes with the help of cholate and incorporating it Consistently, studies with antisera raised against cA and into soybean phospholipid vesicles. In this arti¢cal system, cB from C. heintzii ATCC 29600 indicated the presence of IDA DH activity was fully reconstituted upon addition of cA in both Chelatobacter and Chelatococcus strains, while the major quinones in the Chelatobacter genus as inter- cB was found to be restricted to Chelatobacter strains mediate electron carriers, i.e. ubiquinone Q1 or ubiqui- [6,218]. cA and cB could immunologically only be detected none Q10, and of iodonitrotetrazolium chloride as terminal in cells which were grown with NTA. Interestingly, in electron acceptor [217]. trimethylamine-cultivated C. heintzii bacteria, a protein of 70 kDa, thus signi¢cantly larger than NTA MO, cross- 5.1.3. NTA degradation in denitrifying bacteria reacting with cA antiserum was found. In cell extracts, Obviously, the initial step in the metabolism of NTA in however, no NTA MO activity but only trimethylamine- denitrifying NTA degraders has to proceed via an oxygen- stimulated NADH2 oxidation was detectable [218]. independent step. In membrane-free extracts of strain Also DNA^DNA hybridisation data suggested that the TE11 grown under anoxic conditions, a protein complex gene encoding cA is more conserved than the cB gene [225] consisting of two enzymes, a NTA DH and a nitrate re- (see Section 5.1.4). This con¢rms the crucial role of cA ductase (NtrR), catalysed the ¢rst step of NTA oxidation (NTA MO) for NTA degradation in contrast to that of resulting in the formation of IDA and glyoxylate [228] cB, which probably can be replaced by any other enzyme, (Table 4). Only with both enzymes present, NTA was providing reduced FMN. Here it is noteworthy that ^ transformed, and activity was coupled to a phenazine- apart from the luciferase system and the two-enzyme sys- methosulfate (PMS)-dependent transfer of electrons from tems listed above ^ other NADH2 :FMN oxidoreductases NTA to nitrate which was reduced to nitrite. Apart from providing free reduced £avins have been described which the arti¢cial dye PMS, no naturally occurring compound seem to be involved in the reduction of iron complexed to has been found so far which mediates the electron transfer siderophores [227]. However, no similarity at the amino from NTA DH to NtrR. As for NTA MO, the substrate acid sequence level among the various classes of these spectrum of the DH seems to be restricted to NTA. oxidoreductases has so far been found. N-Ethylmaleimide, a thiol-binding reagent, inhibited An unusual property of the oxidoreductase cB from NTA DH but an excess of dithiothreitol partly restored C. heintzii was that the reduction of FMN was highly enzymatic activity, indicating the involvement of thiol stimulated by the addition of NTA. Other structurally groups in reaction catalysis by NTA DH [229]. As a redox related compounds, among them also IDA, had no such component NTA DH contains a covalently bound FAD stimulating e¡ect when tested in reaction mixtures con- moiety. Additionally, an iron content of about four iron taining both cB and cA [216]. According to these NADH2 atoms per mol of enzyme was found. Since results from oxidation assays, the substrate spectrum of the NTA MO di¡erence spectra of oxidised versus reduced enzyme ar- (cA) was stated to be rather narrow. Yet, catalysis by the gued against the presence of heme chromatophores, NTA NTA MO of the oxidative splitting of some of those com- DH is supposed to contain two 2Fe^2S clusters, one in pounds that did not stimulate the NADH2 oxidation by each monomer of the homodimeric enzyme [229]. Thus, cB cannot be totally excluded. Nevertheless, IDA is cer- electrons derived from NTA are presumably ¢rst trans- tainly not further attacked by the NTA-degrading enzyme ferred to FAD, then to the iron^sulfur cluster, and ¢nally system as originally proposed by Firestone and co-workers to PMS or the so far unknown in vivo electron carrier. [164] since no glyoxylate is formed from IDA in the pres- The in vitro found NTA DH/NtrR complex is rather ence of cA and cB [216]. unusual and in fact the ¢rst enzyme complex reported where a catabolic DH is so tightly associated with a 5.1.2. Enzymatic degradation of IDA NtrR. Therefore, the in vivo localisation of both NTA Since the NTA MO does not catalyse the cleavage of DH and NtrR was investigated by immunochemical label- IDA, an additional enzyme must exist. Indeed, IDA-oxi- ling of the enzymes to address the question of whether the dising activity was detected in the particulate membrane complex is an artefact produced during cell disruption or fraction from cells of C. heintzii and C. asaccharovorans, whether it really exists in the cells. NTA DH was detected which was subsequently attributed to a membrane-bound in the cytoplasm, whereas NtrR was associated with or M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 91 integrated in the cytoplasm membrane [225]. The reduc- of nmoT with several transposases indicate that it may tase from the NTA DH/NtrR complex appears to belong be part of an insertion element. to the dissimilatory type of NtrRs [225] because NtrR To test the presence of genes homologous to ntaA and expression was not observed under conditions where ntaB in other NTA-degrading strains, DNA^DNA hybrid- only the assimilatory type of NtrRs should be produced, isation experiments were performed using di¡erent restric- i.e. during aerobic growth with nitrate as sole nitrogen tion fragments from cloned parts of the cA and cB genes source. In contrast, ammonium did not repress the reduc- from C. heintzii ATCC 29600 as probes [225]. Surpris- tase expression under anoxic conditions. Moreover, NtrR ingly, in C. heintzii TE6, only regions homologous to the production was independent of the presence of NTA in the cA gene and ORF1 but not to the gene for cB were re- growth medium while NTA DH was only formed during vealed although previous immunological studies had indi- cultivation with NTA [225]. cated the presence of proteins similar to cA and cB [218]. NTA DH was also puri¢ed from aerobically grown Conversely, no crossreacting proteins similar to cB were strain TE11, and seemed to be identical to the enzyme detected in C. asaccharovorans TE2, whereas at the DNA expressed under anoxic conditions [229]. However, during level, regions highly homologous to those of ORF1 and aerobic growthm, electrons were transferred to the respi- the genes encoding cA and cB in C. heintzii ATCC 29600 ration chain, but again, only in the presence of PMS. were found. However, shortly after the start of the ntaB Although the NTA DH/NtrR complex from denitrifying equivalent gene from strain TE2, the homology with ntaB cells and NTA DH from aerobically grown cells catalysed from C. heintzii ATCC 29600 ended. Apparently, the gene the same overall reaction, association of NTA DH with for cB in C. asaccharovorans TE2 is disrupted in its read- NtrR led to a marked change in the kinetic properties with ing frame, and most probably no intact cB is synthesised. an increase in a¤nity for NTA [6]. However, it remains to An attempt was made to clone the gene for NTA DH be tested whether this is real or an artefact originating from denitrifying strain TE11 [225] starting with oligonu- from the puri¢cation procedure. cleotide probes derived from the N-terminal amino acid The interesting aspect of the process of in vivo electron sequence of the enzyme [229]. However, because the spe- transfer from NTA DH to either NtrR or to the respira- ci¢city of the available sequence information was too low, tion chain is a still unsolved and intriguing question, es- the probes synthesised hybridised with many di¡erent pecially when bearing in mind that NTA DH apparently is DNA fragments. This is consistent with the observation not associated with the cytoplasmic membrane. In analogy that the N-terminal amino acid sequence of NTA DH to the situation found in case of the functionally related exhibited high similarity to other £avin-containing en- trimethylamine DH [230], an electron-transferring £avo- zymes all possessing a common FAD-binding segment protein was suggested to be involved [229], but this hy- (LKL-fold) in their N-terminus [229]. pothesis has not yet been con¢rmed. Also in the denitrifying strain TE11 ^ similar to the 5.2. Catabolism of EDTA and the similarity of EDTA MO obligately aerobic Chelatobacter and Chelatococcus strains with NTA MO ^ IDA is further oxidised by a membrane-bound IDA DH [229]. Due to the successful isolation of EDTA-degrading bacterial strains within the 1990s, ¢rst information on the 5.1.4. Genetic information concerning NTA metabolism metabolism of EDTA could be obtained. However, so Quite recently, the NTA-metabolising enzymes de- far only strain BNC1 and DSM 9103 have been studied scribed above were also investigated at the genetic level biochemically while the pathway for EDTA breakdown in [219,220,225]. The genes of the two enzymes, cA and cB the Agrobacterium strain growing with Fe(III)EDTA is (ntaA/nmoA and ntaB/nmoB, respectively), were cloned still unknown. from C. heintzii ATCC 29600. They were oriented diver- Uptake of EDTA into cells of strain DSM 9103 was gently with an intergenic region of 307 bp, a rather un- mediated by an energy-dependent carrier, which is most usual organisation for genes whose products act so closely probably driven by the proton motive force. This trans- together. Downstream of the gene for cA, additional open port system had a high apparent a¤nity for EDTA. It was reading frames (ORFs) (ORF1/nmoR and nmoT) were rather speci¢c for EDTA because, of several APCAs and found. In its N-terminus, the 24.4 kDa gene product of other structurally related compounds, only DTPA compet- ORF1/nmoR exhibited a DNA-binding motif (helix-turn- itively inhibited the transport of EDTA [231]. helix) which is characteristic for a family of bacterial reg- Based on products excreted by a mixed microbial cul- ulatory proteins called GntR family. Hence, the ORF1/ ture during EDTA degradation, Belly and co-workers nmoR gene product might be involved in the regulation [143] proposed two possible pathways for the microbial of the expression of the genes for cA and cB. Unfortu- breakdown of EDTA, i.e. either the successive removal nately, the attempts to further characterise the putative of acetyl groups from EDTA or the cleavage of a C^N NTA regulator have failed so far. Sequence similarities bond within the ethylenediamine part of the molecule. The former mechanism, the removal of acetyl groups, was re- 92 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Fig. 5. MO-catalysed degradation of NTA and EDTA as found in NTA degrader C. heintzii and EDTA-degrading strain DSM 9103 [198,216,220]. cently con¢rmed in cell-free extracts obtained from strain propanoltetraacetic acid. N,NP-EDDA was not further BNC1 [193,232] and from strain DSM 9103 [198]. The transformed by EDTA MO but a cofactor-independent EDTA-splitting activity required molecular oxygen, N,NP-EDDA-degrading activity was detected in cell-free NADH2 and FMN and yielded glyoxylate and ^ transi- extracts of strain DSM 9103 [198]. Additionally, N,NP- ently ^ ED3A, the latter being further degraded under the EDDA transformation was also observed in membrane same assay conditions resulting in the formation of anoth- fractions of DSM 9103 cells. By both, the soluble and er glyoxylate and N,NP-EDDA. In both strains, no N,N- the particulate fraction, N,NP-EDDA was oxidatively split EDDA production was observed, indicating regiospeci¢c resulting in the formation of one glyoxylate molecule. Ap- removal of the second acetyl group from EDTA. parently, an additional acetyl residue was removed from From strain DSM 9103, a two-enzyme system was pu- the ethylenediamine backbone probably leaving EDMA ri¢ed which catalysed the removal of two acetyl groups (Fig. 5). Preliminary data indicate the involvement of a from EDTA (Fig. 5, Table 4) [198]. Most likely, one en- DH [233]. However, the enzyme(s) responsible for this zyme (cAP) is a MO catalysing the oxidative cleavage of activity still remain(s) to be isolated. EDTA and ED3A while consuming oxygen and reduced One can speculate that EDMA is further metabolised by FMN (it has not been demonstrated yet that molecular elimination of the last acetyl moiety leaving ethylenedi- oxygen is incorporated into water and glyoxylate). The amine which can be transformed to glycine. Alternatively, second enzyme (cBP)isaNADH2 :FMN oxidoreductase EDMA might be cleaved within the ethylenediamine part that provides FMNH2 for the MO. As in case of NTA of the molecule yielding ¢nally two molecules of glycine. MO, the oxidoreductase from P. ¢scheri could take over As a third possibility, EDMA might be converted into the function of cBP. Furthermore, cB from the NTA-oxi- IDA by the removal of the primary amine group. IDA dising enzyme system was able to replace cBP in the can then be oxidised to form glycine and glyoxylate. The EDTA-degrading enzyme system, and vice versa. metabolic pathway for the products of EDTA degrada- In contrast to the NTA-degrading enzyme system, tion, glycine and glyoxylate, is assumed to be similar to EDTA MO exhibited a broader substrate spectrum. It that described previously for the NTA degrader C. heintzii cleaved not only EDTA and ED3A but also other APCAs [233]. such as NTA, DTPA, HEDTA, PDTA and 1,3-diamino- All in all, important similarities between the NTA me- M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 93 tabolism in C. heintzii and the EDTA breakdown in strain DSM 9103 are discernible. In both cases, a MO catalyses the ¢rst attack on the chelating agent and the MO is dependent on an accessory enzyme providing reduced FMN. Less acetylated intermediates of the metabolism (carrying only two acetyl groups) are then further converted by a DH. The close taxonomical neighbourhood of the organisms as well as these biochemical similarities foster the hypothesis that the enzymes have common ancestors. Only genetic studies, in particular on strain DSM 9103, can give an answer to this question. Just recently, an EDTA MO from bacterium BNC1 was puri¢ed and characterised [234]. In contrast to EDTA MO from strain DSM 9103, this enzyme catalysed the removal of only one acetyl group from EDTA resulting in the formation of ED3A and glyoxylate, whereas formation of EDDA could not be detected. However, whether or not ED3A is accepted as a substrate by the EDTA MO from strain BNC1 was not tested. On the other hand, the EDTA MOs from both strains share the property of act- ing together with a distinct oxidoreductase providing re- Fig. 6. Lyase-catalysed degradation of EDDS in the presence and absence of metal ions. Note that it is not known which metal species is duced FMN. the true substrate for the lyase [235].

5.3. EDDS catabolism enzymatic degradation of AEAA is also stereospeci¢c EDDS is a structural isomer of EDTA and it can also [161]. While a microbial consortium mineralised [S,S]- be used as a growth substrate by the EDTA-degrading EDDS totally, utilisation of [R,S]-EDDS resulted in the strain DSM 9103. Therefore, it was speculated that formation of a dead-end product, identi¢ed as AEAA EDDS might be transported into the cells by the EDTA which was supposedly present in the D-con¢guration. Con- carrier and subsequently be cleaved by EDTA MO. This, sequently, only l-AEAA, which should be formed during however, proved not to be the case [198,231,235]. Whereas [S,S]-EDDS breakdown, can be expected to be further uptake of EDDS by strain DSM 9103 has not yet been metabolised. studied, the initial step in the intracellular EDDS breakdown has been elucidated [235]. The catabolism of EDDS was initiated by a carbon^ 6. Regulation of the APCA degradation and enzyme nitrogen lyase catalysing the non-hydrolytic cleavage of expression the C^N bond between the ethylenediamine part of the molecule and one of the succinyl residues without any 6.1. Regulation of NTA degradation cofactors being required (Fig. 6, Table 4). The reaction led to the formation of fumarate and AEAA. Also in The ability to metabolise NTA was inducible in Chela- C. asaccharovorans, a similar [S,S]-EDDS-splitting activity tobacter, Chelatococcus and denitrifying strains. NTA- and was detected requiring no cofactors and resulting in the IDA-degrading activity was only detected in cell-free ex- formation of the same products as those found in DSM tracts and membrane fractions, respectively, of bacteria 9103, thus indicating the action of the same type of en- grown with either NTA or IDA [100,209]. This, however, zyme [235]. provides only limited information concerning the regula- The further degradation of AEAA remains still to be tion processes taking place under conditions found in unraveled. To date, one can merely speculate that, cata- wastewater treatment plants or in the environment which lysed by a DH or a MO, the C^N bond between the are characterised by rather low substrate concentrations. succinyl residue and the ethylenediamine part of the mol- Therefore, experiments with C. heintzii ATCC 29600 were ecule is split, or that an aspartyl residue is removed by the conducted in carbon-limited continuous culture fed with cleavage of a C^N bond within the ethylenediamine part either NTA or glucose to get a better insight in possible of AEAA [233]. regulation mechanisms of NTA breakdown during growth Out of the three stereoisomers of EDDS ([S,S]-, [R,R]- under conditions similar to those found in the environ- and [R,S]-EDDS), the lyase accepted only [S,S]- and [R,S]- ment [236,237]. EDDS. Probably only the S-con¢guration of the chiral C- In a glucose-limited continuous culture, i.e. in the ab- atom can be attacked by the enzyme. Apparently, the sence of NTA, both the NTA- and IDA-oxidising activ- 94 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 ities were very low and close to the detection limit, inde- and spectrum caused by shock loading, diurnal and weekly pendent of the growth rate applied. When cultivated with cycles, heavy rainfalls, etc. Consequently, cells have to NTA as sole source of carbon and nitrogen, the speci¢c adapt to the changing conditions, and the lag phase result- activity of NTA MO increased with decreasing dilution ing from these adaptation processes can severely a¡ect the rates. The cells exhibited a maximum speci¢c activity (ap- e¤ciency of biodegradation over long time periods [241]. proximately 180 Wmol NTA (g protein)31 min31) at dilu- To investigate the dynamic behaviour of NTA degradation rates between 0.02 and 0.03 h31, whereas at a dilution tion in a continuous culture of C. heintzii under transient rate close to Wmax of C. heintzii, the speci¢c NTA MO conditions, the feed was switched from a medium contain- activity was 4^5 times lower. Thus, at high growth rates ing glucose only to one containing either NTA only or with NTA, repression of the NTA MO activity was ob- mixtures of glucose and NTA [237]. When glucose-pre- served, a pattern often found for enzymes catalysing the grown cells were suddenly confronted with NTA as sole ¢rst step in catabolic pathways [238,239]. Under the same source of carbon, nitrogen and energy, a long lag phase of conditions, IDA DH activity was approximately constant about 25 h was measured until NTA MO expression at growth rates above 0.03 h31 and only decreased con- started. This lag phase was considerably shortened when siderably at lower growth rates. the cells were supplied with mixtures of NTA and glucose In both nature and wastewater treatment plants, NTA- instead of NTA only. Possibly in the latter case, the por- degrading micro-organisms will not only encounter NTA tion of glucose available after the switch was able to sup- as sole substrate but they will grow in the presence of a port the rearrangement of the cellular metabolism, partic- complex mixture of potential carbon and/or nitrogen sour- ularly to provide energy and building blocks for the ces [240]. Indeed, under laboratory batch conditions, both synthesis of NTA-degrading enzymes. Hence one can ex- Chelatobacter and Chelatococcus spp. are capable of con- pect that alternative carbon substrates have a considerable suming NTA in combination with a suitable carbon sub- positive e¡ect on the expression of other enzymes under strate generally resulting in an enhanced speci¢c growth changing environmental conditions. rate [177]. To investigate the degradation of NTA during Although such steady-state and dynamic experiments mixed substrate growth in more detail, C. heintzii was with pure cultures of C. heintzii provide some basic infor- cultivated with di¡erent mixtures of glucose and NTA in mation on the behaviour of this bacterium, one still can- a carbon-limited chemostat and special attention was giv- not infer how NTA degradation is really regulated in en to the expression of enzymes involved in NTA metab- wastewater treatment plants or in surface waters. In par- olism [236]. Synthesis and activity of NTA-degrading en- ticular, the question of whether regulation proceeds at the zymes was controlled by the ratio of substrates rather than level of enrichment of NTA-degrading micro-organisms or their actual concentrations in the feed. During growth rather at the level of induction/repression of the pollutant- with mixtures containing less than 1% of the total carbon speci¢c enzyme systems remains unanswered. Neverthe- as NTA, the amount of cA and cB produced was close to less, a number of observations suggest that for short- the detection limit. Although the speci¢c activities of NTA term £uctuations in the supply with NTA regulation takes MO and IDA DH in cell extracts were very low, the bac- place at the enzyme level [188]. Firstly, when NTA was teria were able to degrade the small amounts of NTA added to activated sludge in laboratory batch cultures, provided. Either the constitutive enzymatic activity was only expression of NTA MO but no enrichment of su¤cient to consume the NTA supplied or an alternative, NTA-degrading bacteria from the genera Chelatobacter currently unknown enzyme system was involved. Feeding and Chelatococcus (assessed with surface antibodies) was a NTA/glucose mixture containing 3.6% of NTA^carbon detected within about 100 h. Also, only a slight rise of triggered induction of NTA enzymes above the constitu- Chelatobacter bacteria was observed after activated sludge tive background level and a further increase of the NTA in porous pot plants that had been fed with synthetic fraction resulted in clearly enhanced activities of both wastewater containing high concentrations of NTA for NTA MO and IDA DH. Assuming that under environ- more than 1 month. Secondly, no enrichment of C. heintzii mental conditions NTA-degrading enzymes are similarly was found in the Swiss wastewater treatment plant on the regulated by the proportion of NTA to the total carbon mountain Saentis which was periodically exposed to un- utilised, no signi¢cant induction of these enzymes can be usually high in£owing NTA concentrations. NTA MO, expected in rivers receiving a high load of treated waste- however, became detectable within a few days after the water where NTA may contribute only to some 0.1^1%. In ratio of NTA^carbon to DOC had increased over 10% wastewater treatment plants, however, NTA might con- in the in£uent, resulting in a markedly higher NTA remov- tribute to some 1^10% of the total utilisable carbon, and al e¤ciency [188]. here, induction of NTA-degrading enzymes can be ex- In contrast to these data [188], in situ hybridisation with pected to play a signi¢cant role. 16S rRNA probes speci¢c for bacteria from the Rhizobia- Obviously in the engineered or natural environment, ceae group revealed a steady raise of Rhizobiaceae in a NTA-degrading bacteria will rarely experience steady-state fed-batch culture of activated sludge daily amended with conditions, but rather changes in substrate concentration 1gl31 NTA [242]. Most of the APCA-degrading strains M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 95 known to date belong to this family of Rhizobiaceae. of the metal ion at the outer side of the cytoplasmic mem- However, cells detected by the Rhizobiaceae probe did brane. Alternatively, the cells could transport only free not react with probes speci¢c for C. heintzii, which might APCA existing in equilibrium with the metal-complexed indicate that important NTA-degrading bacteria from the APCA in the surrounding medium. This question of Rhizobiaceae group are still unknown. Unfortunately, in APCA speciation arises also at the level of enzyme-cata- these studies, measurements of NTA concentrations or lysed breakdown in the cytoplasm. enzyme activities in the culture were lacking to prove As to the properties of the metal^APCA species itself, whether or not the observed enrichment was really linked one is left with the question of whether the thermodynam- to the degradation of NTA. ic stability of a complex, its dissociation kinetics, or its structure governs the uptake into a cell or the enzymatic 6.2. Regulation of degradation of EDTA and EDDS transformation. Several di¤culties are encountered when investigating As for the NTA-degrading enzymes, EDTA MO is also the e¡ect of APCA speciation, in particular in assays inducible in the two EDTA-degrading strains BNC1 and with whole cells. Firstly, metal toxicity has to be consid- DSM 9103. When strain DSM 9103 had been grown with ered. Secondly, the presence of metal ions as well as metal- EDTA or ED3A, the rate of EDTA consumption in cell- complexing ligands on the bacterial surface should be tak- free extract was stimulated 5^10-fold over the constitutive en into account, because both might signi¢cantly a¡ect the level of bacteria cultivated with glycerol. Growth with complexation of the APCA added to the system. However, NTA or IDA, however, did not result in such a stimula- it is di¤cult to assess the in£uence of these surface-bound tion [198]. This partly contrasts with the observations re- ligands or metals. Thirdly, although the speciation of ported for EDTA-degrading strain DSM 9103: compared APCAs at the beginning of an assay can be predicted to EDTA-grown cells, both EDTA transport and EDTA- rather easily, changes in the course of such an experiment oxidising activity were 10 times lower when cells had been are di¤cult to assess. grown with IDA or fumarate, whereas it was similarly high (about 80% of the activity of EDTA-grown cells) 7.1. In whole cells when the strain had been cultivated with NTA or [S,S]- EDDS. This also suggests that transport and catabolic 7.1.1. NTA enzymes are regulated in a co-ordinate manner [198,235]. Investigations of the biodegradability of di¡erent NTA Likewise, [S,S]-EDDS-degrading activity in the two species were ¢rst performed in the 1970s. The systems strains DSM 9103 and C. asaccharovorans was inducible. studied (activated sludge, river water, soil) were rather Lyase-catalysed EDDS transformation was only detected complex ^ with many other ligands and metal ions being in cell extracts of [S,S]-EDDS-grown bacteria but not in present beside the NTA complex ^ and, therefore, no con- fumarate-, EDTA- or NTA-grown cells [235]. clusive information was obtained. Nevertheless, the data suggested that NTA complexed with Ca2,Fe3,Mn2 or Pb2 is probably easily biodegradable under environmen- 7. In£uence of metal speciation on microbial APCA tal conditions whilst NTA bound to Cd2,Cu2,Hg2 or degradation Ni2 might be more recalcitrant [132,243^246]. Somewhat later, degradation of metal^NTA complexes In the previous chapters on microbial and enzymatic by a `Pseudomonas' species (now C. heintzii ATCC 29600) APCA degradation, one important point was not touched, was studied under more de¢ned conditions [247]. How- namely the fact that in both growth media and the envi- ever, also here bu¡er systems were used which themselves ronment APCAs are generally complexed with metal ions. have complexing properties and, therefore, the speciation Consequently, metal ions must be expected to have a sub- of NTA was not precisely de¢ned. When metal concentra- stantial in£uence on the biodegradability of APCAs. In tions were kept low enough to avoid toxicity, the bacterial this context, several aspects have to be considered: ini- strain was able to attack Ca2,Mn2,Mg2,Cu2,Zn2, tially, the APCA molecule has to be transported across Cd2 and Fe3 chelates of NTA, as well as free NTA as 14 the cytoplasm membrane metal-free or in association determined by oxygen consumption and CO2 evolution with a metal ion. If the latter were the case, a cell would from 14C-labelled NTA. Only NiNTA was resistant to- encounter an enormous in£ux of metal ions. To deal with wards degradation. this situation, the cell can either excrete the metal ions or Recently, more elaborate data on the e¡ects of NTA inactivate the cations by precipitating them intracellularly speciation on microbial degradation were presented in with a suitable anion (e.g. phosphate). On the other hand, which calculations of the chemical speciation were in- if only the free APCA molecule enters the cell, then the cluded [248]. Degradation assays with C. heintzii ATCC metal has to dissociate from the complexing agent prior to 29600 revealed ¢rst order kinetics for the decomposition of uptake. This could be achieved by destabilising the metal^ several degradable NTA species at concentrations ranging APCA complex during transport, resulting in the release from 0.05 to 5.23 WM NTA with degradation rates in the 96 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

3 3 3 following order: H2NTA s CoNTA VFeOHNTA V Fe(III)EDTA. Free EDTA as well as complexes with Ni, ZnNTA3 s AlOHNTA3 s CuNTA3 s NiNTA3. This Cu or Co(III) did not support growth [190,249]. On the order indicates no relationship with the chemical stability other hand, the EDTA-degrading strain BNC1 was not of the di¡erent complexes. However, a relationship be- capable of utilising Fe(III)EDTA [194]. Growth of this tween the rates of degradation and the lability of the var- strain on EDTA mineral salt medium always stopped be- ious metal^NTA complexes was established. This implies fore the substrate was completely degraded and the re- 3 that the rates of formation of H2NTA from MeNTA, i.e. maining concentration of EDTA corresponded to the the dissociation rates of the complexes, determine the deg- Fe(III) concentration in the medium. Nevertheless, growth radation rates. Two models can explain these results: the of BNC1 was dependent on the presence of metal-com- 3 ¢rst model is based on the assumption that only H2NTA plexed EDTA because free EDTA appeared to interact is taken up into the cells, with the extracellular dissocia- negatively with the cell walls and completely inhibited bac- tion of the metal^NTA complex being the rate-limiting terial growth. step in the degradation. The second model assumes that Experiments with resting cells of strain BNC1 revealed 3 all species, i.e. H2NTA and metal^NTA complexes, enter that EDTA complexes with low stability constants (log the cytoplasm, where the species accepted by the NTA K 6 14) were readily degraded. CaEDTA, BaEDTA and MO has then to be generated. Previously, Uetz and co- MgEDTA were consumed at the highest rates, followed by workers [216] had demonstrated that the preferred sub- MnEDTA. Free EDTA was also degraded, although strate of NTA MO is MgNTA, whereas CaNTA or growth with this species was observed to slow down. Gen- 3 14 H2NTA are not accepted by the enzyme. Therefore, an erally, chelates with stability constants exceeding 10 were exchange of Mg2 for the metal complexed to NTA must not degraded. Only ZnEDTA (log K = 16.5) was slowly take place intracellularly before NTA degradation can oxidised [232]. Similar results were reported recently for proceed. This exchange reaction will be the rate-limiting strain DSM 9103 [250]. Complexes with stability constants step which is again dependent on the dissociation kinetics (log K 6 16, i.e. the Mg2,Ca2,Mn2 complexes and free of the metal complexes. Unfortunately, the data collected EDTA) were degraded by EDTA-grown resting cells to are not su¤cient to establish which of the processes ac- completion at a constant rate. For more stable EDTA tually occurs. For this, short-term studies of NTA uptake chelates (with Co2,Cu2,Zn2 and Pb2), the data sug- into cells, preferably carried out with membrane vesicles, gested that these complexes were not used directly but had are required. Such data are still lacking and experience to dissociate prior to degradation. Only CdEDTA and from our laboratory indicates that it will be very di¤cult Fe(III)EDTA were not degraded within 48 h. In the case to prepare membrane vesicles from C. heintzii [214]. of CdEDTA, the toxicity of freed Cd2 ions most likely Somewhat contrasting results were reported by Madsen prevented a signi¢cant degradation of EDTA, whereas in and Alexander [139]. Based on speciation models, de¢ned the case of Fe(III)EDTA, a combination of metal toxicity media were prepared whose composition favoured the and the very slow dissociation of the complex was prob- presence of one predominant species of NTA. Under ably responsible for the absence of degradation [250]. such conditions, an unde¢ned microbial consortium min- For transport in strain DSM 9103, uptake experiments eralised CaNTA but not AlNTA, MgNTA, Fe(III)NTA with [14C]EDTA in presence of various metal ions revealed or free NTA. The same pattern was found for a Listeria a nearly identical pattern to that described for resting cells sp. isolated in the course of the study although the isola- of strain BNC1 [231]. Free EDTA as well as chelates of tion medium contained mainly free NTA. Con¢rmatory Ca2,Ba2 and Mg2 were readily incorporated at similar tests with this strain in the presence of constant NTA rates. Uptake of MnEDTA proceeded at an intermediate and varying Ca concentrations (9 NTA concentration) rate and all other, more stable complexes (those with showed that the extent of NTA mineralisation increased Zn2,Co2,Cu2,Ni2 or Fe3) were not transported with increasing Ca concentrations. in these short-term assays. In addition, Ca2 uptake of strain DSM 9103 was considerably increased in the pres- 7.1.2. EDTA ence of EDTA, indicating that Ca2 is probably trans- Several reports on the degradability of various EDTA ported together with EDTA into the cells. Cells of strain complexes su¡er from the same drawbacks as those de- DSM 9103 seem to utilise both proposed mechanisms, i.e. scribed above for NTA, i.e. that unde¢ned, rather complex intracellular precipitation and excretion of the metal, to systems were used. In such systems, EDTA was found to cope with the increased incoming £ux of metal ions due be degraded by mixed microbial consortia when added in to EDTA degradation. the free form or as chelate of Cu2,Cd2,Zn2,Ni2, All these studies indicate that for the two strains BNC1 Mn2,Ca2 or Fe3 [149,153]. and DSM 9103, the EDTA chelates of Zn2,Cu2,Ni2, Contrasting requirements of pure EDTA-degrading cul- Co2 and Fe3 are not directly accessible. They can only tures with respect to EDTA complexes supporting growth be degraded after dissociation, a process that depends on were reported. On the one hand, growth of an Agrobacte- the dissociation rate constant of the complex. However, rium sp. was only observed when EDTA was supplied as this rate constant is rather low for some of the heavy M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 97 metal^EDTA complexes, such as Fe(III)EDTA, NiEDTA formed at a very low rate. All other complexes which or CuEDTA [91]. Consequently, the environmental recal- were not attacked by the lyase exhibited stability constants citrance of EDTA towards biological degradation might ^ higher than 1010. Hence, in the case of EDDS, the stability at least partly ^ be attributed to the presence of species of a complex apparently rules the enzymatic degradability. which do not favour microbial attack, in particular when One can speculate that during substrate-binding, the metal considering that in a typical river system about 70% of the ion has to be exchanged or removed from the complex total EDTA is predicted to be present as stable heavy with the EDDS molecule. This assumption is supported metal complexes (Fig. 3) [85]. by the observation that metal ions had a strong in£uence on the equilibrium between EDDS and the cleavage prod- 7.2. At the enzyme level ucts which is reached in the lyase-catalysed elimination reaction (Fig. 6). In presence of metal ions (especially At the enzymatic level, NTA MO from C. heintzii those forming non-degradable complexes), this equilibri- ATCC 29600 [216,251], NTA DH from strain TE11 um was shifted towards the educt EDDS probably because [229,252], EDTA MO from strain DSM 9103 [198], cell- free EDDS was removed from the equilibrium by the for- free extracts and EDTA MO from strain BNC1 [232,234], mation of metal complexes, indicating that only free and ¢nally EDDS lyase from strain DSM 9103 [235] were EDDS is the actual substrate of the lyase. used to test the e¡ects of APCA speciation (see Table 4). With respect to the metal complexes that were accepted as substrates, both NTA MO from C. heintzii and EDTA 8. Outlook MO from strain DSM 9103 exhibited a similar behaviour. Neither of the MOs transformed the free complexing agent To date, the industrially most important APCAs are the or the Ca2 chelate, whereas complexes of NTA or EDTA synthetically produced compounds NTA, EDTA, DTPA with Mg2,Mn2,Co2 and Zn2 were degraded. NTA and HEDTA. Recently, a number of new APCAs, such as MO attacked also Fe(III)NTA and NiNTA. The highest PDTA, L-ADA, SDA or MGDA, have been synthesised speci¢c transformation activity was found for the Mg2 which may have the potential to substitute the classical complexes. Since Mg2 is the most abundant intracellular representatives in various applications and they are cur- divalent cation [253], NTA and EDTA can be expected to rently being tested. In addition, several naturally occurring be mainly present as Mg2 chelate in the cytoplasm. APCAs produced by micro-organisms or plants, also with Therefore, it is not surprising that Mg complexes of considerable potential for application, have been described both APCAs were the best substrates. All in all, no clear in the literature. The producing organisms exploit the met- relationship is discernible between the stability of a NTA al-complexing capacity of APCAs for enhancing their met- or EDTA complex or its dissociation rate constant and its al uptake, in most cases that of iron. On top of their good MO-catalysed oxidation. Therefore, it was suggested that complexing ability, a considerable advantage of these nat- the structure of a complex might be a determining factor ural APCAs is that they are probably better biodegradable [198]. than most synthetic APCAs. This is because, ¢rstly, the It should be added that for the EDTA-degrading strain metal they are transporting into the cell has to be freed BNC1, con£icting results were published. In cell-free ex- intracellularly and, secondly, because organisms have been tracts, Klu«ner and co-workers [232] found a similar pat- exposed to them for a longer time than to the xenobiotic tern to that described above for the two MOs, with free ones. However, to date, there is little information available EDTA and CaEDTA not being oxidised. In contrast, on these aspects. First studies with respect to a biotechno- these two species, as well as various other EDTA com- logical production of some of these APCAs have already plexes with di- and trivalent cations, were degraded in been reported and they seem promising. Nevertheless, con- enzyme assays with puri¢ed EDTA MO performed by siderably more research is still needed before such biotech- Payne and co-workers [234]. Currently, this discrepancy nologically produced APCAs may compete with the estab- cannot be explained. lished and cheap synthetic APCAs NTA and EDTA. In contrast to NTA MO and EDTA MO from DSM Once released into the environment, the fate of APCAs 9103, NTA DH accepted free NTA. Addition of divalent is determined either by abiotic or biotic processes (Fig. 7). metal ions almost completely inhibited enzymatic NTA For each of the di¡erent compounds, it has to be assessed breakdown [252]. This inhibition was reversed upon addi- individually because no general pattern applicable to all of tion of EDTA, strongly suggesting that only uncomplexed them has emerged. The environmental fate of NTA is NTA is a substrate of the DH. determined by biodegradation with abiotic processes play- Similarly, EDDS lyase accepted free [S,S]-EDDS as a ing a minor role for its elimination from both wastewater substrate but also EDDS complexes with Mg2,Ca2, and ecosystems. Hence, NTA most likely does not present Ba2 and Mn2 [235]. All these complexes have stability a major risk for the environment. In contrast, EDTA, constants lower than 109. MnEDDS with the highest DTPA and HEDTA seem to be ^ if at all ^ only poorly stability constant of all degradable species was trans- biodegradable and therefore abiotic elimination mecha- 98 M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Fig. 7. Abiotic and biotic processes determining the fate of APCAs in sewage treatment plants and in the environment. (1) Adsorption of APCAs to positively charged minerals such as iron- or aluminium(hydr)oxides and subsequent sedimentation of the APCA-loaded particles. This process might be of some importance for the elimination of APCAs from the waterbody of lakes and their transport into sediments. (2) Biodegradation of APCAs. Apart from NTA, [S,S]-EDDS and EDTA, also other APCAs were reported to be biodegradable, however, the microbial metabolism of these compounds has not been investigated yet. While biodegradation of NTA is the key mechanism of elimination of this compound from environmental systems, the signi¢- cance of biodegradation for the environmental fate of other APCAs cannot be estimated yet with the data available. (3) Photodegradation of APCAs. This process is only relevant for Fe(III)-complexed APCAs or for free APCAs adsorbed to iron(hydr)oxides. Photodegradation does not result in total mineralisation of the compounds. In the case of EDTA, however, the photodegradation products formed seem to be better biodegradable than the original chelating agent, whereas for other APCAs such as DTPA this question still has to be investigated. (4) Oxidation of APCAs by manganese(III/IV)- oxides. Similar to photodegradation, also this abiotic oxidation does not lead to total mineralisation of the chelating agents but only to a partial deacy- lation. The importance of this process in natural systems for elimination of APCAs is still unknown.

nisms are of greater signi¢cance. Indeed, for EDTA it has potential to replace the classical ones. Some of them ap- been demonstrated that the most important elimination pear to be easily biodegradable, at least under aerobic process in rivers is its photolysis, which is, however, re- conditions. Though, for all of them, the information avail- stricted to sunny days and to only one EDTA species, i.e. able is still too limited to conclude that they are harmless Fe(III)EDTA. To what extent abiotic oxidative processes from an environmental point of view. are important for EDTA (and perhaps also other APCAs) So far, the isolation of aerobic strains growing with breakdown in soils, especially in comparison to biodegra- NTA, EDTA or [S,S]-EDDS has been successful. Virtu- dation, remains to be investigated. Furthermore, one ally all strains have been found to be members of the should not forget that abiotic processes usually do not Agrobacterium^Rhizobium group in the K-branch of Pro- lead to a mineralisation of the complexing agents, but teobacteria. A possible future isolation and taxonomic merely to a transformation into compounds which still characterisation of bacterial strains growing with DTPA, have metal-complexing properties. Therefore, future re- HEDTA, L-ADA, SDA, ASDA or MGDA might unravel search should also focus on the metabolites of these abio- an interesting picture with respect to the evolution of mi- tic degradation processes such as ED3A. Especially in crobial degradation of xenobiotic APCAs. In the absence monitoring programs for EDTA and DTPA potential of molecular oxygen, NTA is the only APCA for which breakdown products should be included, too, for a better degradation has been reported under denitrifying condi- assessment of the pollution of a system with these xeno- tions. Interestingly, this strain is not a member of the biotic APCAs. As mentioned above, new APCAs of both K-branch of Proteobacteria. synthetic and biological origin are being tested for their Some of the NTA- and EDTA-degrading strains pres- M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 99 ently isolated seem to exhibit an extended spectrum of are really taken up by the cells. This information, how- APCAs they can utilise as growth substrates. One can ever, seems essential for understanding why the biodegrad- therefore expect that testing the newly marketed chelating ability of di¡erent APCAs varies so tremendously. If only agents such as PDTA, L-ADA, EDDS or MGDA for free APCA molecules or rather unstable species are trans- degradation by NTA- or EDTA-utilising bacteria might ported into the cells, then this might explain why APCAs yield interesting results. Furthermore, the fact that both such as EDTA or DTPA forming very stable complexes the EDTA-degrading strain investigated in our laboratory (especially with heavy metals) are not as well biodegrad- and the NTA-degrading Chelatococcus strains are able to able as NTA. This could also explain the positive e¡ect of degrade not only one but two or more APCAs indicates a pH rise on the microbial elimination of EDTA as it was that such pure cultures could perhaps be used for special observed in a sewage treatment plant because it is known treatment of wastewaters which contain high amounts of that heavy metal^EDTA complexes are destabilised at rather poorly biodegradable APCAs such as DTPA, higher pH values. HEDTA or PDTA. In case of EDTA, the application of All in all, the enormous in£uence speciation has on the pure cultures has already shown rather promising results fate of APCAs in nature shows clearly the necessity to in laboratory experiments [212,213], whereas large scale include all important species of new APCAs in degrada- treatment still remains to be investigated. tion and toxicity tests in order to be really able to predict Due to the isolation of pure bacterial cultures, the met- and judge the compound's environmental behaviour. abolic pathways for NTA, EDTA and [S,S]-EDDS degradation have been ^ at least partly ^ elucidated. The ¢rst transformation of all three compounds consists of the Acknowledgements cleavage of the bond between a nitrogen atom and a carboxylic residue (an acetyl or succinyl group). So far, de- The authors would like to thank Kay Fox, Laura Sigg carboxylation or cleavage within the ethylenediamine part and Bernd Nowack for careful reading of and valuable of the EDTA or EDDS molecule has not been observed as comments on this manuscript. TE would like to express ¢rst metabolic step. In this context, it is worth mentioning his thanks to Hansueli Weilenmann for the fruitful collab- that the biochemistry of EDTA degradation in the Agro- oration over the years (not only in the ¢eld of complexing bacterium strain which grows only with iron(III)-com- agent biodegradation) as well as to all the research stu- plexed EDTA [190], a complex that is not touched by dents that have added mosaic stone by mosaic stone with- the strains investigated so far, has not been studied yet. out which it would have been impossible to get to the Perhaps, the co-ordinated iron ion plays a role in the present level of understanding. Furthermore, the generous enzymatic breakdown of EDTA in this bacterium. ¢nancial support from the Swiss National Science Foun- Of course, for the characterisation of degradative path- dation, The Research Commission of the Swiss Federal ways for other APCAs such as DTPA or the amino acid Institute of Technology Zu«rich, Lever Switzerland and derivatives L-ADA, SDA, ASDA or MGDA, pure cul- Unilever Reserach Laboratories Port Sunlight (UK), the tures able to degrade these chemicals should be available. Swiss Agency for the Protection of the Environment However, one can speculate that the aerobic breakdown (BUWAL), and last, but not least, of the Swiss Federal will also include the attack of an MO on the compounds Institute for Environmental Science and Technology is as seen for NTA and EDTA degradation in strains gratefully acknowledged. C. heintzii, BNC1 and DSM 9103. This would lead to the removal of an acetyl group. 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