Environmental fate and microbial degradation of aminopolycarboxylic acids

Share Embed


Descripción

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 * ë berlandstrasse 133, CH-8600 Du«bendorf, Switzerland Swiss Federal Institute for Environmental Science and Technology, Department of Microbiology, U 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 NTAutilising 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,NEDDA 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: S 0 1 6 8 - 6 4 4 5 ( 0 0 ) 0 0 0 5 5 - 3

FEMSRE 706 29-12-00

70

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Contents 1.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

70 70 73 76 77 78 79 80 80 80 81 82 82 83 84 84 84 84 85 86 87 87 88 91 93 93 93 95 95 95 97 97

Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

99

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

99

2.

3.

4.

5.

6.

7.

8.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1. Aminopolycarboxylic acids (APCAs) ^ an important group of chelating agents . . . 1.2. Important APCAs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3. Fields of application for APCAs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4. Concentrations of APCAs in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5. Speciation of APCAs in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.6. Environmental risks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Abiotic elimination of APCAs in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. NTA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. EDTA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3. Other APCAs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Microbial degradation of APCAs during wastewater treatment and in the environment . 3.1. NTA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. EDTA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. DTPA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4. Other synthetic APCAs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5. Naturally occurring APCAs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Degradation of APCAs by pure microbial cultures . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1. Pure cultures utilising NTA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. Pure cultures utilising EDTA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3. Potential of bacterial isolates for the treatment of APCA-containing wastewaters . . Biochemistry and genetics of APCA degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1. NTA degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2. Catabolism of EDTA and the similarity of EDTA MO with NTA MO . . . . . . . . . 5.3. EDDS catabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Regulation of the APCA degradation and enzyme expression . . . . . . . . . . . . . . . . . . . . 6.1. Regulation of NTA degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2. Regulation of degradation of EDTA and EDDS . . . . . . . . . . . . . . . . . . . . . . . . . . In£uence of metal speciation on microbial APCA degradation . . . . . . . . . . . . . . . . . . . 7.1. In whole cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2. At the enzyme level . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Outlook . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction 1.1. Aminopolycarboxylic acids (APCAs) ^ an important group of chelating agents APCAs, also known as complexones, are compounds that contain several carboxylate groups bound to one or more nitrogen atoms (Fig. 1). Hence, they are essentially derived from the amino acid glycine [1,2]. The major chemical property of APCAs is their ability to form stable and water-soluble complexes with many metal ions. APCAs form a special kind of metal complexes because they co-ordinate the metal ion by forming one or more heteroatomic rings (Fig. 2). This ring formation leads to a higher stability of the complexes as compared to metal^ligand complexes in which no such rings are present, a phenomenon which is called `chelate e¡ect'.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Additionally, the presence of basic secondary or tertiary amino groups and the large negative charge of APCAs contributes to the high stability of metal^APCA complexes. To describe the stability of such a metal^ligand complex, the equilibrium constant K is used. It is de¢ned as Kˆ

‰MeLŠ ‰MeŠW‰LŠ

…1†

where [Me] is the concentration of the metal ion, [L] the concentration of the ligand, and [MeL] that of the ligand^ metal complex at equilibrium [2]. In many industrial processes and products, the presence of free metal ions causes problems such as the formation of insoluble metal salt precipitates or the catalysis of the decomposition of organic compounds. Chelation masks the metal ions and restricts them from playing their nor-

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

mal chemical role and from entering potentially harmful and unwanted reactions. Because of their industrial importance, chelating agents are produced and used in large quantities. Among the chelating agents presently employed, polyphosphates are the most widely used compounds, followed by APCAs [3]. Therefore, much attention has been paid to the environmental fate of APCAs, with interest focused on nitrilotriacetate (NTA) and ethylenediaminetetraacetate (EDTA), the two most important APCAs. In the case of NTA, research concentrated on biodegradation because this was found to be the major environmental elimination pathway for this compound. Several reviews have already been dedicated to this topic [4^6]. For EDTA, however, biodegradation does not seem to be a relevant elimination mechanism and so far an abiotic mechanism has been identi¢ed as major process for the degradation of EDTA in natural systems [7]. However, an increasing number of reports on the biodegradation of EDTA have been published recently and also other APCAs are receiving more and more attention concerning their occurrence and behaviour in the environment. In this review, we will summarise the information concerning the microbial breakdown of APCAs. At the same time, we would like to present the most important aspects of the abiotic degradation of this group of compounds to provide a current summary of the knowledge of the fate of APCAs in the environment. 1.2. Important APCAs The predominant representatives of APCAs are the synthetically produced compounds EDTA, NTA, diethylenetriaminepentaacetate (DTPA) and hydroxyethylethylenediaminetriacetate (HEDTA). Figures on the total amounts produced of the most important synthetic APCAs are di¤cult to obtain and the estimated amounts of these chelating agents produced or used in the USA and Western Europe are listed in Table 1. Their chemical

Table 1 Estimated use of synthetic chelating APCAs (in 103 metric tons) in the USA and Western Europe in 1981

EDTA NTA DTPA HEDTA Other APCAs Sodiumpolyphosphates Organophosphonates Hydroxycarboxylic acids

USA

Western Europe

42 32 4 18 2.5 586 10 110

13.6 8.3 0.5 2.0 2.0 1111 10 15

For comparison, data are also shown for other industrially important complexing agents. Figures for the USA are based on production, those for Europe are estimated from information on consumption (adapted from [3]).

71

structure is shown in Fig. 1A. In recent years, an increasing number of APCAs of biological origin have been described, most of which are involved in metal acquisition by organisms (Fig. 1B). 1.2.1. Synthetically produced APCAs In 1862, the ¢rst chemical synthesis of an APCA, namely that of NTA, was described [8]. It was based on the reaction of monochloroacetic acid in an ammonical solution. Much later, in 1935, the synthesis of EDTA was reported by I.G. Farbenindustrie. Again, the synthesis was based on monochloroacetic acid reacting with ethylenediamine in the presence of sodium hydroxide. Alternatively, EDTA is produced from ethylenediamine reacting with sodium cyanide and formaldehyde in the presence of sodium hydroxide. Depending on the amine employed, also other APCAs, for instance 1,3-propylenediaminetetraacetate (PDTA), can be produced by this type of reaction [2]. The most important function of APCAs is to form metal chelates. This chelation has a strong in£uence on their environmental fate, including biodegradation. We will therefore summarise the properties of the most intensively investigated APCA chelates, i.e. the metal complexes of NTA and EDTA. NTA contains four donor atoms and is a so-called quadridentate chelating ligand. It forms 1:1 complexes with metal ions by establishing three chelate rings with four co-ordination sites of the metal occupied by NTA. Since most cations have a co-ordination number of six, the remaining two sites are normally occupied by water molecules, resulting in octahedral co-ordination of the ion (Fig. 2A) [1]. EDTA contains six donor atoms and acts as a hexadentate ligand. It can form a maximum of ¢ve chelate rings. Ideally, EDTA should form an octahedral complex in which both the metal ion and EDTA have a co-ordination number of six (Fig. 2A). However, this octahedral co-ordination seems to be only possible with cations of relatively 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 accessible to other ligands such as water molecules. Indeed, Xray analysis has shown that the structure of most metal EDTA complexes di¡ers from the ideal octahedral structure and that the cations exhibit often higher co-ordination numbers than six (Fig. 2B). On the other hand, in some complexes, such as those with Cu2‡ or Ni2‡ , EDTA does not fully utilise its donor capacity by leaving one carboxylate group not co-ordinated. Octahedral co-ordination is instead completed by a water molecule (Fig. 2B) [1,2]. Stability constants for 1:1 complexes of EDTA and NTA are listed in Table 2. For each metal, the stability of its NTA complex is several orders of magnitude lower than that of its EDTA complex, as can be expected from

FEMSRE 706 29-12-00

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 Fe(III)EDTA has a stability constant of log K = 25.0, this is still not high enough to keep it from decomposing at pH values above 8^9 due to precipitation of iron(III)hydroxide. In the search for more e¡ective chelating agents at higher pH values, DTPA and HEDTA were developed. They are preferentially employed instead of EDTA for sequestering iron(III) ions in the pH range of 8^10. EDDHA with a log K of 33 (Fig. 1A) is able to complex iron(III) ions more selectively than EDTA,

DTPA or HEDTA and it does not decompose even in the most strongly alkaline solutions [2]. Lately, the synthesis of several new APCAs was reported, namely L-alaninediacetate (L-ADA), serinediacetate (SDA), asparaginic acid diacetate (ASDA) and methylglycinediacetate (MGDA) (Fig. 1). They are derived from amino acids by substituting the amino group with two acetyl groups. The stability of their metal complexes seems to be closer to NTA complexes than to EDTA^ metal complexes [9]. Also, the synthesis of N-acyl deriva-

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

73

Fig. 1 (continued).

tives of ethylenediaminetriacetate (ED3A), chemicals that act both as a surfactant and powerful chelating agent, given the short name chelactant, has been described [10].

1.2.2. Naturally occurring APCAs Recently, it was shown that APCAs are also naturally occurring compounds. Some even contain an ethylenediamine central moiety, a feature that was considered to

FEMSRE 706 29-12-00

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. 1.2.2.1. Ethylenediaminedisuccinate (EDDS). The ¢rst natural APCA described, EDDS, was isolated from culture ¢ltrate of the actinomycete Amycolatopsis orientalis. It was detected in an antibiotic screening program due to its ability to inhibit activity of the Zn2‡ -dependent phospholipase C [11]. In the actinomycete, EDDS is most probably involved in Zn2‡ uptake, because its production was found to be totally repressed at Zn2‡ concentrations higher than 2.5 WM in the growth medium [12], whereas below 2.5 WM it increased with decreasing Zn2‡ concentrations. In contrast, other trace metals (Fe2‡ , Fe3‡ , Co2‡ , Mn2‡ , Mo2‡ ) exerted no repressive e¡ect on EDDS production [13]. However, there is no preferential complexation of Zn2‡ by EDDS, for the stability constants for Fe3‡ , Co2‡ , Ni2‡ and Cu2‡ are higher than that for Zn2‡ (Table

Table 2 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 M at a temperature of 25³C or ^ where indicated (+) ^ at 20³C (¢gures were taken from [254]) Metal ion 2‡

Mg Ca2‡ Mn2‡ Zn2‡ Co2‡ Cu2‡ Pb2‡ Cd2‡ Al3‡ Fe2‡ Fe3‡ Ni2‡

FEMSRE 706 29-12-00

Log KMeNTA

Log KMeEDTA

Log KMeEDDS

5.47 6.39 7.46 10.66 10.38 12.94 11.34 9.78 11.4 8.33 (+) 15.9 11.5

8.83 10.61 13.81 16.44 16.26 18.7 17.88 16.36 16.5 14.27 25.0 18.52

5.82 4.23 8.95 (+) 13.49 (+) 14.06 18.36 12.7 (+) 10.8 (+)

22.0 (+) 16.79

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

2) [14]. Also, it has not been demonstrated yet that Zn2‡ is really transported by EDDS into cells of A. orientalis. EDDS exhibits two chiral C-atoms resulting in the existence of three di¡erent stereoisomers, [S,S]-, [R,S]- and [R,R]-EDDS. A. orientalis exclusively produces the S,Sisomer. For the biosynthesis of [S,S]-EDDS, two pathways were proposed by Cebulla [12], starting from either L-aspartate and serine, or from oxaloacetate and 2,3-diaminopropionic acid. In both cases, N-(2-aminoethyl) aspartic acid (AEAA) would be an intermediate reacting then with oxaloacetate to form [S,S]-EDDS. EDTA is rather recalcitrant towards biological and chemical degradation in the environment and there is increasing pressure to replace it by other powerful complexing agents. For this, EDDS seems to be a promising candidate. Consequently, its chemical and biotechnological production has been investigated. Two routes exist to synthesise EDDS chemically: ¢rstly, the reaction of maleic anhydride with ethylenediamine yielding a mixture of the three stereoisomers of EDDS [15,16], and secondly, the reaction of aspartic acid with 1,2-dibromoethane leading to the formation of either pure [S,S]-EDDS or [R,R]EDDS, depending on the stereoisomer of the aspartic acid employed [17]. As an alternative, both the fermentation conditions for the production of [S,S]-EDDS by A. orientalis and the subsequent puri¢cation procedure were optimised [18]. Highest [S,S]-EDDS productivity was obtained by fed-batch cultivation with a medium containing very low Zn2‡ concentrations, glycerol as carbon source, a mixture of glutamate and urea as nitrogen source, and high initial phosphate concentrations. Under optimised conditions, ¢nal EDDS concentrations of some 20 g l31 were obtained. EDDS can be puri¢ed using a three-step procedure consisting of an acid precipitation, an ethanol washing step and a ¢nal crystallisation leading to 92% purity with an overall yield close to 50% [18]. 1.2.2.2. Rhizobactin. In 1984, Smith and Neiland [19] described the formation of the APCA rhizobactin by Rhizobium meliloti strain DM4 when cultured on a low-iron medium. Consequently, rhizobactin is considered a siderophore. As EDDS, rhizobactin contains an ethylenediamine group with one amine function linked to pyruvic acid, and the second to the carbon skeleton of lysine carrying malic acid on the N6 amino group [20]. The absolute con¢guration of rhizobactin was Dala , Llys , Lmalate , suggesting that rhizobactin is biochemically related to the pyruvic acid-derived opines which have a con¢guration of Dala , amino acid L . 1.2.2.3. Rhizoferrin. Also among fungal siderophores, a representative of APCAs, rhizoferrin, was detected. Rhizoferrin was ¢rst isolated from cultures of Rhizopus microsporus var. rhizopodiformis [21]. Apparently, rhizoferrin is a widespread siderophore within the class of Zygomycetes [22]. It is composed of two citric acid groups linked via

75

amid bonds to a 1,4-diaminobutane (putrescine) molecule [23]. The asymmetric carbon atoms of the citric acid residues are always in R,R-con¢guration. A value of 1023:5 was reported for the stability constant of the iron(III) rhizoferrin complex [24]. Thus, rhizoferrin is not a particular powerful iron chelator at pH 7.4 compared to other fungal and bacterial siderophores of the hydroxamate or catecholate type but it becomes much more competitive at pH values near 4^5, the growth optimum for the producing fungi [24]. Transport experiments with 55 Fe-labelled rhizoferrin con¢rmed that the compound acted as siderophore [21]. In the producing fungal strain, Absidia spinosa, reductively inert gallium rhizoferrin, and the kinetically inert chromium and rhodium complexes were also taken up, indicating that the intact metal^siderophore complex was transported into the cell [25]. Rhizoferrin is able to deliver iron not only to the producing organisms but to other micro-organisms as well. It showed strong siderophore activity in nearly all bacterial strains tested from the Proteus^Providencia^Morganella group, although these strains possess their own e¤cient iron uptake system based on keto- or hydroxycarboxylate iron complexes [26]. The rates of iron uptake in Morganella morganii mediated by rhizoferrin were equivalent to those measured in the presence of some of its own siderophores. Recently, a rhizoferrin receptor from M. morganii has been cloned [27]. This transport system (called Rum, for rhizoferrin uptake into Morganella spp.) was speci¢c for ferric rhizoferrin and staphyloferrin A (described below), while citrate and the citrate derivative aerobactin were not transported. The system consisted of two proteins, RumA located in the outer membrane and the periplasmic RumB. Homology analyses indicated a close relationship of RumA to the outer membrane receptor FecA of the Escherichia coli ferric citrate transport system. RumB, however, showed little homology to other binding proteins with the exception of certain signatures of conserved amino acid residues characteristic of binding proteins of the siderophore transport family. The biotechnological production of rhizoferrin and some of its analogues was also studied [28]. From several producing strains tested, Cunninghamella elegans and Mucor rouxii exhibited highest productivity and were selected for process optimisation. Moreover, C. elegans was used to produce analogues of rhizoferrin by feeding analogues of the two building blocks (1,4-diaminobutane and citric acid). Feeding citric acid analogues resulted in only low amounts of derivatives being formed. However, feeding of 1,5-diaminopentane or 1,3-diaminopropane caused a considerable formation of analogues containing bridges with ¢ve or three methylene residues, respectively. Also, bis-(2aminoethyl)-ether and branched diamine precursors were used as building blocks but only low product yields were obtained. Finally, two interesting e¡ects were observed when feeding the ketone 1,4-diamino-2-butanone as a pre-

FEMSRE 706 29-12-00

76

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

cursor. Firstly, the synthesis of rhizoferrin itself was dramatically reduced; and secondly, the analogue 2-oxorhizoferrin was formed preferentially, which was never observed when other 1,4-diaminobutane analogues were supplied. This probably results from 1,4-diamino-2-butanone being a competitive inhibitor of ornithine decarboxylase, an enzyme known to be responsible for the synthesis of 1,4diaminobutane in fungi, bacteria and plants. Hence, in C. elegans, diaminobutane seems to be a precursor of rhizoferrin and ornithine decarboxylase is probably responsible for diaminobutane synthesis. Furthermore, feeding Lornithine led to a preferential excretion of rhizoferrin whereas addition of D-ornithine resulted in a signi¢cant reduction of rhizoferrin produced. All these biologically produced rhizoferrin analogues had siderophore activity in growth promotion assays with M. morganii with activities comparable to or only slightly lower than that of rhizoferrin. In contrast, the citric acid analogues exhibited signi¢cantly lower activity. These results show clearly that also the structure of biologically produced APCAs can be manipulated, a point which might become interesting with respect to a possible future exploitation of naturally occurring APCAs. 1.2.2.4. Staphyloferrins. Structurally related to rhizoferrin is staphyloferrin A, a siderophore originally isolated from a culture of Staphylococcus hyicus [29,30]. It consists of two citric acid residues linked by two amid bonds to Dornithine [30]. As in the case of rhizoferrin, the tying together of two citric acid residues provides an iron-binding agent much more powerful than citrate alone [29]. Staphylococci produce also another siderophore of the complexone type, staphyloferrin B, the chemical structure of which is quite di¡erent from that of staphyloferrin A. Composed of 2,3-diaminopropionic acid, citrate, ethylenediamine and 2-ketoglutaric acid as structural components, it lacks the symmetry of staphyloferrin A and rhizoferrin, and it is therefore a less potent chelator of iron [26]. Both staphyloferrin siderophores are produced by a wide range of Staphylococcus strains, some of which were even found to produce both types [31]. Transport measurements in S. hyicus with ferric staphyloferrin A and B revealed uptake of 55 Fe within the range of 1^15 min with staphyloferrin B being less e¡ective than staphyloferrin A [26]. 1.2.2.5. Plant siderophores. Siderophores from plants were described which also ¢t into the group of APCAs. An example is nicotianamine, an essential constituent of higher plants which is important for cellular iron transport and/or metabolism. Nicotianamine contains six donor groups (three nitrogen atoms and three carboxylate groups) with an arrangement ideal for the formation of chelate rings [32]. Presently, there is no evidence that nicotianamine plays a role in iron uptake from the rooting media, whereas this is known to be the case for the chemically similar mugineic acid and its acidic amino acid ana-

logues, as well as avenic acid (Fig. 1B). These phytosiderophores are apparently widespread among grass species and cultivars [33]. Mugineic acids might also be produced and excreted by plants stressed by zinc de¢ciency [34]. Mass production of mugineic acid and its derivatives is proposed because of their potential application in agriculture, medical science and pharmacy. Possible production methods take advantage of cell cultures of barley or wheat, or are based on water cultures of intact plants, in both cases exposing the cells or plants to iron de¢cient conditions [35]. 1.3. Fields of application for APCAs Due to their metal sequestering capacity, synthetic APCAs are used in many industrial processes or products in order to (i) prevent the formation of metal precipitates, (ii) to hinder metal ion catalysis of unwanted chemical reactions, (iii) to remove metal ions from systems, or (iv) to make metal ions more available by keeping them in solution. 1.3.1. Use of APCAs to prevent formation of precipitates In addition to active washing ingredients, detergents contain a large proportion of metal-complexing agents to inhibit the formation of insoluble Ca2‡ and Mg2‡ salts, and thus prevent the deposition of scale on both textile ¢bres and washing machine parts. The ¢rst complexing agents employed in modern detergents were di- and triphosphates [36]. However, it was soon found that these contribute to the eutrophication of lakes and rivers. During the search for substitutes, NTA was proposed as an alternative detergent builder and NTA-containing detergents were ¢rst marketed in Sweden in 1967. Since then, NTA-based detergents have been used in Canada, Finland and Sweden, the US, and Switzerland, although they have met considerable opposition [4] (note that EDTA is not used as a substitute for polyphosphates in household detergents, see below). In industrial cleaning agents, NTA, ETDA, and recently also MGDA, are used to prevent precipitation of calcium, magnesium and heavy metal salts [37^39]. For the same purpose, EDTA is especially used in the textile and photographic industry, as well as in electroplating processes instead of cyanide [37]. In the photographic industry, FeNH4 EDTA is also employed as oxidising agent for silver [37], but recently it is being partly replaced by L-ADA and PDTA [39]. 1.3.2. Use of APCAs to prevent catalysis mediated by metal ions Small amounts of EDTA have been included in many detergent formulations to stabilise the bleaching agent perborate by preventing metal-catalysed decomposition of the compound [38]. In the bleaching process of the pulp and paper industry, hydrogen peroxide is increasingly employed instead of chlorine compounds and the addition

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

of either EDTA or DTPA avoids the decomposition of hydrogen peroxide catalysed by manganese or iron ions [40]. APCAs are also used as additives for pharmaceuticals, cosmetics and food to prevent transformation of the ingredients or rancidity due to metal-catalysed reactions [37,41]. 1.3.3. Use of APCAs to remove metal ions Multidentate chelating agents are widely applied in the nuclear industry for decontamination of reactors and equipment because they form water-soluble complexes with many radionuclides. Although the exact composition of most of the decontamination reagents is proprietary information, it is assumed that mainly NTA, EDTA, HEDTA or DTPA are present in these formulations [42]. In a single decontamination operation, hundreds of kg of chelating agents are employed [42,43]. The wastes obtained, containing radionuclide^chelator complexes, are solidi¢ed and disposed of [42^44]. Concern was raised that due to the presence of powerful chelators in such wastes the migration of radioactive transition metals, rare earths and transuranics may be enhanced [45]. APCAs have also been proposed for application in the remediation of metal-contaminated soils or sediments, either to serve as washing agents [46,47] or to support electrokinetic extraction processes [48,49]. Apparently, APCAs have also the potential to be employed in phytoremediation strategies due to their ability to increase metal desorption from soil and to facilitate metal uptake by plants [50]. Moreover, in medical treatment, CaEDTA is used as antidote for the treatment of lead or other heavy metal intoxications. EDTA complexes the toxic metal and accelerates its excretion [51]. 1.3.4. Use of APCAs to increase metal availability Since the early 1950s, synthetic chelating agents have been used to improve plant nutrition [52]. Especially APCA chelators are employed in fertilisers to supply plants with trace metals such as iron, copper, zinc and manganese. Most commonly, the Fe(III) chelates of EDTA, HEDTA, DTPA, EDDHA and ethylenediaminedi(o-hydroxy-p-methylphenyl) acetic acid and the Cu(II), Zn(II) and Mn(II) chelates of EDTA are present in fertilisers [53]. 1.4. Concentrations of APCAs in the environment Since the 1970s, NTA has received much attention because of its use in laundry detergents and this has led to monitoring programs during which its concentration was measured in various environmental compartments. More recently, interest was also focused on EDTA and the determination of EDTA concentrations was often included in the monitoring of NTA. In contrast, little is known about the occurrence and concentrations of other APCAs in the environment.

77

1.4.1. Concentrations of NTA and EDTA in sewage treatment plants In Canada, where NTA accounts for up to 15 volume % of laundry detergents, the typical NTA loading of raw wastewater averaged 2500 Wg l31 [54]. Several case studies in Swiss municipal wastewater treatment plants revealed similar average concentrations in the in£uent of 100^1000 Wg NTA l31 , whereas EDTA concentrations in raw wastewater were determined to be somewhat lower ranging from 10 to 500 Wg EDTA l31 [55^57]. However, the EDTA concentrations in secondary e¥uents were approximately ¢ve times higher than those of NTA, and did not di¡er signi¢cantly from the EDTA concentrations in the in£uent [57]. Obviously, a large portion of the incoming NTA, but not of EDTA, was eliminated from the wastewater during the treatment process. 1.4.2. Concentrations of NTA and EDTA in surface waters Despite the dramatic increase of NTA usage in Canada since 1970, no accumulation of NTA in Canadian surface waters was observed [58]. From 1971 until 1975, a median concentration of 50 Wg NTA l31 was found in Canadian streams [59]. In Canadian coastal waters of the Atlantic and the Paci¢c Ocean and in samples from the port area of Halifax, no NTA was detectable more than 2 years after its ¢rst usage in detergents [60]. The concentrations of NTA and EDTA found in European rivers are in the range of 0^20 and 0^60 Wg l31 , respectively [37,38,61]. For EDTA, however, occasionally concentrations higher than 100 Wg l31 were found in German and Swiss rivers [62,63]. In Swiss lakes, NTA concentrations measured were below 10 Wg l31 . At the bottom of the lakes, NTA was present at approximately the detection limit of 0.1^ 0.2 Wg l31 . EDTA concentrations of 1^4 Wg l31 were found throughout the whole waterbody and in contrast to NTA, the concentrations of EDTA were only slightly subject to temporal or spatial variations [61]. In several sediment cores from Lake Greifensee, Switzerland, EDTA concentrations in the range of 60 Wg kg31 to 1170 Wg kg31 were found [64]. The water content of the sediments was 80% with an EDTA concentration in the pore water equal to that found in the overlaying water column, namely around 6 Wg l31 , indicating an enrichment of EDTA in the sediments of this lake. A similar enrichment of EDTA in the sediment, ranging between 80 and 310 Wg EDTA kg31 , was found in the Southern part of the Finish Lake Saimaa [65]. This lake receives a high load of treated wastewaters from pulp and paper industries, which use EDTA and DTPA in their production process [65, 66]. 1.4.3. Concentrations of NTA and EDTA in ground- and drinking water NTA concentrations determined in ground- and drinking water ranged normally between 1 and 5 Wg l31

FEMSRE 706 29-12-00

78

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

[58,61,67]. In Swiss groundwaters, EDTA concentrations of 0.1 to nearly 15 Wg l31 were found [61]. Drinking water obtained from surface water by bank in¢ltration from the river Ruhr showed rather high EDTA concentrations (median of 25 Wg l31 ), similar to those found in the supplying river (median of 26 Wg l31 ). This indicates that no elimination took place in the water along the in¢ltration passage. For NTA, however, the concentrations decreased by a factor of 10 during in¢ltration, resulting in a median concentration of 0.7 Wg l31 in the drinking water [67]. Due to the observed persistence of EDTA during in¢ltration, the compound was proposed to be used as tracer to analyse how far in¢ltrated river water penetrates into groundwater aquifers [68]. 1.4.4. Environmental concentrations of DTPA Only little data are available on concentrations of DTPA in freshwater. In the few samples collected from German rivers, DTPA concentrations between 2 and 15 Wg l31 were determined [38]. Rather high concentrations for both DTPA and EDTA (10^20 Wg l31 ) were observed in the Southern part of the Finnish Lake Saimaa. In contrast to EDTA, which was found in all samples, DTPA was only detected in the vicinity of the point sources and it was not found in the sediments despite its presence in the overlaying waterbody [65,66]. Even though in these investigations the analytical detection limit for DTPA was higher than that for EDTA and the extractability from the sediments might have been lower, this suggests that DTPA was more readily degraded in this particular lake water than EDTA. 1.4.5. Environmental concentrations of other APCAs The occasional samples originating from various German rivers revealed the presence of L-ADA and PDTA at concentrations around 1^5 Wg l31 . In the same samples, MGDA was not detectable. Furthermore, HEDTA was measured at two sampling points in the river Elbe (Germany) at rather high concentrations of about 40 Wg l31 [39,69]. These preliminary data indicate that it is not suf¢cient to analyse only the `classical' representatives NTA, EDTA and DTPA but that also other complexing compounds have to be considered when monitoring the pollution of surface waters with APCAs. 1.4.6. Concentrations of APCAs in soils We are not aware of reports on the concentration of APCAs in agriculturally used soils despite their usage in fertilisers. Eventually, the highest environmental APCA concentrations were reported for the Hanford site, USA, where wastes from nuclear decontamination processes containing both complexing agents and radioactive metals were dumped into the ground [70]. In the mixed waste, approximately 9 g EDTA l31 , 10 g HEDTA l31 and 1^2 g NTA l31 were detected plus a wide variety of chelator fragments, also present in the mg^g l31 range.

1.5. Speciation of APCAs in the environment In addition to environmental concentrations, information on the speciation of APCAs is essential to understand their environmental fate. So far, only data on the speciation of NTA and EDTA have been reported. Considering typical concentrations of metal ions and APCAs in natural systems, the chelating agents can be expected to occur mostly as metal complexes. Unfortunately, reliable analytical methods for monitoring the various metal species of APCAs in natural waters are still lacking. Yet, in the case of EDTA, a method based on high performance liquid chromatography has been developed and ¢rst promising results have been obtained. However, this method is still merely applicable to high concentrations of the various EDTA species [71]. Attempts to exploit capillary electrophoresis have been made to assess the speciation of APCAs, in particular of EDTA and DTPA. But this method is restricted to rather stable complexes and its sensitivity is still too low for detecting complexes in environmental samples [72,73]. Nonetheless, making use of the photolability of Fe(III)EDTA and the slow exchange kinetics of NiEDTA with Fe(III) ions, analytical methods were developed to distinguish Fe(III)EDTA and NiEDTA from other EDTA complexes even when present at low, environmentally relevant concentrations [7,64]. Generally, equilibrium calculations based on the stability constants (see Eq. 1) have to be used to predict the speciation of APCAs in ecosystems. However, two major di¤culties a¡ect the prediction of APCA speciation under environmental conditions. Firstly, other natural ligands including inorganic anions such as hydroxide, bicarbonate and phosphate anions and organic compounds such as humic acids compete with APCAs for the metals. Whereas concentrations of inorganic ligands can be measured and thus their e¡ect on APCA speciation can be predicted, it is more di¤cult to estimate the in£uence of organic ligands on metal complexation. For marine systems, methods to analyse the complexation of a variety of metal ions with natural ligands were described [74^77]. For freshwater systems, such methods exist for Cu2‡ and Zn2‡ [78,79] and Co2‡ and Cd2‡ [80,81]. The second di¤culty is that speciation calculations assume that the chemical equilibrium has been reached. However, and this applies in particular to EDTA, due to the slow kinetics of some of the metal exchange reactions, true speciation in a natural system may di¡er considerably from the calculated equilibrium. Under conditions as they are found in natural waters, where a large excess of Ca2‡ and Mg2‡ over trace metals exists, exchange reactions of EDTA complexes have been shown to occur at slow rates with time scales of hours to days [82,83]. Especially Fe(III)EDTA was observed to exchange rather slowly with other metals and for the exchange reactions with divalent cations, half-life times of about 20 days were determined in river water [84]. Hence,

FEMSRE 706 29-12-00

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

Fig. 3. Speciation of EDTA and NTA in the Swiss river Glatt as predicted by speciation calculations and, in case of Fe(III)EDTA and NiEDTA determined analytically. Data taken from [85].

the complexation of EDTA in river water depends not only on the dissolved concentrations of the various cations and other ligands which determine the equilibrium speciation but also on the initial speciation of EDTA in which it is released. This explains why Fe(III)EDTA makes up a considerable portion of total EDTA in surface waters, because it is mainly this species which is released into the environment from both photographic industry and wastewater treatment plants using iron salts for phosphate precipitation. Several authors presented speciation calculations for NTA or EDTA, taking into account the above mentioned problems to a larger or smaller extent. Early predictions for NTA speciation in freshwater including both inorganic ligands and humic acids forcasted the chelating agent to be mostly bound to Cu2‡ when present at environmentally relevant concentrations of 10^20 Wg l31 [51]. According to more recent speciation modelling for a small river, which receives an extensive load of treated wastewaters, NTA present at 2 Wg l31 should be predominantly complexed with Ca2‡ (more than 90% of the total NTA) followed by Mg2‡ and Zn2‡ and only a small portion of NTA would be associated with other heavy metals (Fig. 3) [85]. The inclusion of natural ligands for Cu2‡ and Zn2‡ into these speciation predictions might explain the discrepancy with the earlier calculations. Apparently, the natural ligands for Cu2‡ compete successfully with NTA for the metal ion. The contradicting modelling data underline the need for more information on the e¡ect of natural ligands on metal speciation. To determine the speciation pattern of EDTA in a river system, a combination of measurements and equilibrium calculations was applied (Fig. 3). According to measurements, the Fe(III)EDTA and NiEDTA fractions were set at 31% and 10%, respectively. Including into the model

The main concerns which have been raised in the discussion 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 potential to mobilise metals. 1. No negative e¡ects on the normal operation of wastewater treatment plants or sludge disposal systems have been reported so far for both EDTA and NTA at environmentally relevant concentrations [4,37]. 2. Acute and chronic toxicity of NTA towards more than 50 species of freshwater and marine organisms has been studied [4]. Chronic `no observed e¡ect concentrations' of NTA for aquatic life were at least one order of magnitude higher than measured environmental concentrations [4]. Furthermore, acute or chronic e¡ects were only reported when the NTA concentrations used in the tests were equal to or in excess of the concentration of divalent metal ions. In analogy, also EDTA and DTPA were only weakly to moderately toxic for aquatic organisms when they were complexed with metal ions [37,87,88]. Given the fact that in surface waters always a large stoichiometric excess of Ca and other divalent metal ions is present and that the actual NTA, EDTA and DTPA concentrations are many orders of magnitude below the observed toxic concentrations, no adverse e¡ects on aquatic life can be anticipated for these APCAs. In toxicity studies, NTA was only moderately toxic to mammals during acute oral exposure. It was not teratogenic itself or in presence of heavy metals and it was non-genotoxic. NTA was not metabolised but was rapidly excreted by the kidney. Findings that urinary tract tumours can develop as a consequence of chronic exposure to high doses of NTA were explained by changes in Zn and Ca distribution between urinary tract tissues and urine. But again, thresholds determined for such e¡ects were much higher than human exposure resulting from the low environmental NTA concentrations [4]. Also EDTA is only weakly toxic to man (in fact, it is used as antidote and is permitted as food additive in some countries). WHO ¢xed the acceptable daily intake of EDTA to 2.5 mg kg31

FEMSRE 706 29-12-00

80

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

body weight. Considering average concentrations of EDTA in drinking water, the daily uptake of EDTA is far below this value [37]. 3. It was suspected that widespread usage of NTA or EDTA (both contain nitrogen) might enhance eutrophication. However, the contribution of both compounds to the environmental nitrogen loading was found to be insigni¢cant for any direct eutrophication e¡ect [4]. It has also been discussed whether or not NTA or EDTA would stimulate algal growth indirectly by extracting essential metals from sludges, sediments or humic acid and making them better available or ^ alternatively ^ by protecting organisms against the toxic e¡ects of certain heavy metals. While in laboratory systems indirect e¡ects were demonstrated at relatively high NTA and EDTA concentrations, they are thought to be negligible in surface waters [4,37]. 4. The potential of NTA and EDTA to alter heavy metal distribution has been thoroughly studied in model systems (for reviews, see [4,37,89]). It was concluded that metal mobilisation by NTA and EDTA will not be signi¢cant at environmental concentrations and that variations in pH will have larger e¡ects on aqueous concentrations of heavy metal ions. However, shock loadings in wastewater treatment plants with concentrations of free NTA or EDTA in the mg l31 range can cause mobilisation of Zn2‡ leading to elevated Zn2‡ concentrations in the e¥uent [55,90]. Moreover, at pH 7^8 the mobilisation of toxic heavy metals (e.g. Pb or Cd) sorbed to particles such as iron(hydr)oxides resulted from an exchange of Fe(III) complexed by EDTA against the toxic metal, a reaction which is catalysed by the particle surface [91]. Thus, Fe(III)EDTA, which is an important EDTA species in many rivers, can remobilise adsorbed heavy metals during in¢ltration or in groundwater aquifers in calcareous regions characterised by pH values s 7.0. In fact, such unwanted exchange reactions have already been observed in a ¢eld investigation of an in¢ltration passage [91]. On the other hand, in aquifers with lower pH values ( 6 7.0), Me(II)EDTA chelates will react with iron(hydr)oxides resulting in the dissolution of the mineral and the formation of Fe(III)EDTA. The metal liberated from the original EDTA complex will then adsorb to the surface and thus become immobilised. Hence, under such conditions, no remobilisation but rather an immobilisation of toxic metals is to be anticipated [92].

elimination has to be considered separately for each compound. Numerous laboratory and ¢eld investigations have shown that biodegradation is the key mechanism for the removal of NTA from the environment [4], whereas EDTA is predominantly eliminated via photodegradation [62]. Before focusing on the biodegradation of APCAs, we will therefore summarise the most important abiotic processes leading to either transformation of APCAs or to their elimination from environmental compartments.

2. Abiotic elimination of APCAs in the environment

2.2.1. Photolysis The process considered to be the most important for the elimination of EDTA from surface waters is direct photolysis, which results from the fraction of sunlight below 400 nm [62]. Apparently, only Fe(III)EDTA is susceptible to sunlight irradiation, whereas other environmentally rel-

The elimination of various APCAs from the environment is based on di¡erent biotic and abiotic processes. Despite the chemical and structural similarity of the APCAs, the mechanism primarily responsible for their

2.1. NTA For both Fe(III)NTA and CuNTA, photochemical degradation by sunlight was reported, and this may ^ although only to a minor extent ^ contribute to the decomposition of NTA in the photic zones of lakes and marine systems [4,93,94]. Photodegradation of other metal complexes is not likely, since no signi¢cant decrease in NTA concentrations was measured when solutions containing NTA and an excess of Cd2‡ , Pb2‡ , Mg2‡ or Cr3‡ were exposed to light of 350 nm, a wavelength at which both CuNTA and Fe(III)NTA were degraded. The half-life for Fe(III)NTA during irradiation with sunlight was determined to be approximately 1.5 h, whereas that for CuNTA was more than 100 times higher. The decomposition probably originates from a ligand to metal charge transfer resulting in a reduced metal ion and the formation of a ligand radical which then undergoes sequential decarboxylation giving rise to CO2 and formaldehyde, and iminodiacetate (IDA). Fe(III)IDA and Cu(II)IDA were further photodegraded with glycine being formed. The latter, however, degraded more slowly than Fe(III)IDA, because of a shift of the absorption spectrum of Cu(II)IDA towards shorter wavelengths and thus a smaller overlap with the solar emission spectrum [95]. 2.2. EDTA The recalcitrance of EDTA towards biodegradation in wastewater treatment plants or the environment has directed much attention to other mechanisms of elimination. Thermic hydrolysis and indirect photolysis of EDTA are obviously negligible for its fate in natural systems [7]. Direct photodegradation, oxidation by metal(hydr)oxides and, to a smaller degree, also sorption of EDTA to particles and subsequent sedimentation of these EDTAloaded particles seem to be important processes for the partial elimination of EDTA from aquatic systems.

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

evant EDTA species (complexes with Mg2‡ , Ca2‡ , Ni2‡ , Cu2‡ , Zn2‡ , Cd2‡ and Hg2‡ ) will not photolyse. Under laboratory conditions, Mn(II)EDTA and Co(III)EDTA also photodecomposed, although at rates approximately one order of magnitude lower than that of Fe(III)EDTA [96,97]. Interestingly, initially uncomplexed EDTA was photodegraded in the presence of lepidocrocite, an iron oxide [98], indicating that EDTA adsorbed to the surface of the iron hydroxide can be photooxidised. Although free EDTA is not present in natural systems, the authors suggested that metal complexes might undergo photodegradation in an analogous manner by adsorbing to iron oxides and forming ternary surface complexes. Several researchers [40,99,100] have calculated the photolysis half-lives of Fe(III)EDTA for surface waters at various geographical locations. They ranged between only 11.3 min to more than 100 h depending on the light conditions employed for modelling. A comparison of ¢eld data with model calculations demonstrated that photolysis by sunlight is really the most important process for the degradation of EDTA in the Swiss river Glatt [7]. While on cloudy days no signi¢cant decrease in the EDTA concentrations could be detected along the river, all available Fe(III)EDTA was eliminated by photodegradation under sunny conditions within approximately 1 day. The remaining EDTA found on sunny days consisted of photostable EDTA complexes. As products of the photolysis of Fe(III)EDTA, ED3A, both possible isomers of ethylenediaminediacetate (EDDA) (N,N-EDDA and N,NP-EDDA), IDA and ethylenediaminemonoacetate (EDMA) were found, suggesting that the reaction mechanisms are similar to those already described for the photolysis of NTA. Photolysis clearly does not result in the total mineralisation of EDTA but more easily biodegradable metabolites are formed during this process [101]. After a rather short irradiation (6.5 h) of a Fe(III)EDTA solution with sunlight, almost all of the initial EDTA was transformed into ED3A, EDDA and EDMA. When this mixture of photolysis metabolites was incubated with activated sludge, a bioelimination of 53% was observed within 4 weeks. Using a Fe(III)EDTA solution irradiated for 20 h which contained mainly EDDA and EDMA, bioelimination of the photodegradation products was 92%. These results indicate a higher resistance of ED3A towards biological mineralisation when compared with EDDA and EDMA. Clearly, only investigations with pure compounds can give a ¢nal answer. If ED3A actually turns out to be rather persistent, then further ¢eld investigations should be considered which do not only include the monitoring of the disappearance of Fe(III)EDTA but also the fate of ED3A. 2.2.2. Sorption Several studies reported negligible adsorption of EDTA on humic acids, silica, kaolin, river sediments, humus solids and activated sludge particles [55,56,102]. Recently, the

81

adsorbed EDTA fraction was quanti¢ed in some Swiss surface waters and wastewater treatment plants and it ranged around 1% of total EDTA, with 1036 ^1035 g EDTA adsorbed to 1 g of particles [64]. Because of the negative charges of metal^EDTA complexes at pH values typically found in natural waters (between pH 5.0 and pH 8.0), the complexes sorb only to solids that are positively charged, such as aluminium(hydr)oxides, iron(III)(hydr)oxides and manganese(III,IV)oxides [7,103,104]. Whereas adsorption onto such oxides does not play a role in the elimination of EDTA in wastewater treatment plants and rivers, it could do so in lakes due to the sedimentation of the particle-bound EDTA fraction [105]. This hypothesis is supported by the detection of enhanced concentrations of EDTA in sediments compared to the water column and pore water ([64] ; see also Fig. 7). 2.2.3. Redox processes Lately, attention was drawn to the possible oxidation of APCAs by +III metal oxidants in natural systems [106,107]. Such oxidants comprise manganese(III/IV)(hydr)oxides, Co(III)-containing mineral phases (e.g. Co(III) stemming from the oxidation of Co(II) adsorbed onto manganese(III/IV)- and iron(III)(hydr)oxides) as well as iron(III)(hydr)oxides. While the potential of the Fe(III)/ Fe(II) half reaction (E³ = +0.67 V) is too low to allow oxidation of EDTA, those for the Mn(III)/Mn(II) and Co(III)/Co(II) half reactions (E³ = +1.50 V and E³ = +1.48 V, respectively) are su¤ciently high. Indeed, interactions of EDTA with Mn(III)OOH and Co(III)OOH have been shown to result in the oxidation of EDTA, yielding the deacetylated breakdown products ED3A and EDDA. Mn(III) is a more reactive oxidant than Co(III) and, therefore, the reactions of EDTA with Mn(III)OOH were signi¢cantly faster than those with Co(III)OOH. Apparently, EDTA does not persist in the presence of manganese(III/IV)(hydr)oxides and reactions with such minerals were proposed to represent an important sink for APCAs in the environment [107]. However, most experiments were done with free EDTA, and the in£uence of metal complexation of the APCA molecule on the oxidative process still has to be investigated in detail. 2.3. Other APCAs Photodegradation was also reported for Fe(III)DTPA but the products of the reaction were not determined [40]. A theoretical half-life of Fe(III)DTPA lower than that of Fe(III)EDTA was calculated. Hence, the rates of photodegradation of the iron(III) chelates apparently increase in the order NTA 6 EDTA 6 DTPA. DTPA as well as HEDTA also adsorb to surfaces with pH-dependent charge, i.e. aluminium- or iron oxides [108]. Yet, we are not aware of data on the adsorption of DTPA and HEDTA in aquatic systems and their role for the elimination of these APCAs from the waterbody.

FEMSRE 706 29-12-00

82

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

3. Microbial degradation of APCAs during wastewater treatment and in the environment It has been demonstrated in numerous studies that NTA is easily biodegradable during wastewater treatment, in natural waters and in soils. In addition, many investigations indicate that biodegradation is complete and that IDA, a metabolite, which is formed intracellularly during the microbial breakdown of NTA, does not accumulate in nature [4]. In contrast to NTA, contradictory results have been published concerning the biodegradation of EDTA or DTPA in the environment and only few reports can be found in the literature dealing with the biodegradability of other APCAs. Hence, the question of the importance of biodegradation for the environmental fate of these chelating agents is still unanswered. 3.1. NTA 3.1.1. During wastewater treatment During the treatment of wastewater, NTA is mainly eliminated in biological treatment steps such as oxidation ponds and lagoons, activated sludge systems, or trickling ¢lters. Systems that previously had not been exposed to NTA frequently require an adaptation period of 1^4 weeks before maximum degradation activity is achieved [4]. Elimination e¤ciencies for NTA reported for di¡erent wastewater treatment plants range between 70 and s 90%. Less e¤cient NTA removal (in some cases merely around 50%) was only found at low temperatures during the winter months or in treatment plants characterised by elevated sludge loading rates [54,109^115]. NTA elimination during wastewater treatment was successfully simulated with a simple activated sludge model [116] which included only temperature-dependent growth of NTA-degrading microorganisms with NTA [WNTA (T) = 2.2 day31 exp(0.07 (T320³C))], but the model neglected decay of NTA degraders and their growth with carbon sources other than NTA (the two e¡ects probably compensating for each other). Due to lack of sorption, NTA is accumulated neither in primary nor in secondary sludge and, therefore, the removal of NTA during anaerobic sewage sludge digestion is not of utmost importance. Nevertheless, Klein [117] and Kirk and co-workers [118] demonstrated NTA reductions in digesters varying from 8 to 45%, elimination of NTA close to 100% was observed when activated sludge already acclimated to NTA was added to the digestion process [119]. 3.1.2. In the aquatic environment According to numerous laboratory studies, NTA is biologically degraded in freshwater [4,120]. Contradictory results have been reported on its degradation in marine and estuarine water samples. Whereas Kirk and co-workers [121] found no NTA degradation by a marine bacterial

population in coastal marine waters under aerobic or anoxic conditions, several other researchers found NTA degradation in estuarine and o¡shore environments [122^124]. In particular, in samples from an estuarine system which had been pre-exposed to NTA for several years, NTA was rapidly degraded with no lag period at the low NTA concentration of 10 Wg l31 [123]. Apparently, NTA degradation follows ¢rst order kinetics, especially at the low NTA concentrations (below 2 mg NTA l31 ) found in surface and groundwaters [125]. Similar rate constants (about 0.3 day31 ) were determined for all freshwater systems situated in areas where NTA was not extensively used in detergents at the time of the study. They were approximately 3^4 times lower than those found in a river from a region where NTA-containing detergents were marketed. This again underlines the crucial role of acclimation of the microbial population present [123,126]. Similar to activated sludge from sewage treatment plants a few days to 4 weeks were typically required for an unadapted microbial population to acquire NTA-degrading ability. Acclimation to NTA and subsequent biodegradation occurred even at low environmental concentrations between 5 and 50 Wg NTA l31 [126], but the time period needed for developing degradation ability seemed to decrease with increasing NTA concentrations, as tested for a concentration range from 0.02 mg NTA l31 to 20 mg NTA l31 [127]. Analysis of measured NTA concentrations from a ¢eld study along a Swiss river receiving a high load of treated wastewater indicated biological NTA elimination with a half-life of about 8 h [62,128]. This agrees with the results of the laboratory studies described above [126]. Also the e¡ect of temperature on NTA degradation was investigated [62,126,129]. While in summer about 90% of NTA was eliminated from the river Glatt over a £ow distance of 22 km, only 65% of the NTA disappeared in winter [62]. Similarly, model calculations suggested that the behaviour of NTA in a Swiss lake can be attributed to biological degradation as only removal process with a constant degradation rate of 0.035 day31 , corresponding to a half-life of 20 days [105]. 3.1.3. In aquifers and soils In laboratory column systems with aquifer material from a natural river water/groundwater in¢ltration site, NTA was rapidly mineralised under both aerobic and denitrifying conditions [130]. According to these data, it was expected that very low NTA concentrations ( 6 2 Wg l31 ) will be found in the groundwater even in cases where the in¢ltration distance was only a few meters. Hence, even a signi¢cant increase in NTA concentration in river waters should not lead to higher concentrations in the groundwater. Low NTA concentrations measured in various groundwater samples, also in those obtained from in¢ltration areas in close distance to the feeding river, con¢rmed these predictions [38,61,67,131].

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Also in soils, NTA readily decomposed under aerobic conditions with half-lives ranging from 3 to 7 days [132^ 134]. In contrast to Tiedje and Mason [132], both Tabatabai and Bremner [135] and Ward [133] observed NTA degradation also under anaerobic conditions where nitrate substituted molecular oxygen as a terminal electron acceptor. While nitrate concentrations in the system had no in£uence on the initial rates of NTA degradation, they a¡ected the extent of the mineralisation under anaerobic conditions [133]. 3.2. EDTA 3.2.1. During wastewater treatment and in laboratory test systems Balances of the EDTA load in the in£uent and e¥uents of municipal wastewater treatment plants with sampling periods of several days gave no indication for a signi¢cant elimination of EDTA neither by biological nor by physicochemical processes [55,56,136,137]. At the same time, NTA was e¤ciently eliminated ( s 85%) in all plants investigated. Also, several laboratory studies with inocula collected from industrial or municipal wastewater treatment systems consistently demonstrated the recalcitrance of EDTA [138^141]. However, e¤cient EDTA elimination of about 80% in an industrial wastewater treatment plant has been reported lately [142]. In activated sludge samples taken from the very plant, almost 100% of added EDTA (measured as DOC) was degraded within 10 days. Several reports have already described biologically mediated EDTA degradation under laboratory conditions. The ¢rst report [143] demonstrated decomposition of EDTA by microbial populations from an aerated lagoon receiving EDTA-containing industrial e¥uents. The authors followed [14 C]CO2 formation from the iron(III) complex of radioactively labelled EDTA when incubated in the dark to prevent photodegradation. After an incubation period of 5 days, about 90% of the initially present EDTA had disappeared. 27% of the initial radioactivity of the acetate-labelled and 31% of the ethylene-labelled EDTA was recovered as 14 CO2 , indicating that both the ethylene backbone and the acetyl groups were attacked. During EDTA degradation, the concentrations of ED3A and IDA increased transiently. Most likely, these two compounds were intermediates of the microbial EDTA degradation. In contrast, other possible intermediates such as NTA, N,N-EDDA and N,NP-EDDA, EDMA and glycine were found only at very low concentrations. Supplementing the culture with NTA or ethylenediamine resulted in a stimulation of EDTA degradation whereas the addition of amino acids and sugars had an inhibitory e¡ect. Unfortunately, the study was not extended to the fate of EDTA in the lagoon from where the culture was obtained to give an indication whether or not the consortium was also consuming EDTA under environmental conditions.

83

Later, Gschwind [144], working under aerobic conditions with micro-organisms ¢xed on carrier materials, observed degradation of EDTA leading to the formation of nitrate (probably because of the presence of nitri¢ers). With this system, treatment of a model wastewater containing 200 mg EDTA l31 under continuous conditions led to 99% elimination at a hydraulic retention time of 1.5 h. Addition of yeast extract or glucose did not a¡ect the performance. Interestingly, the optimum pH for e¡ective EDTA degradation was rather high (between pH 9.0 and 9.5). A similarly high optimum pH for EDTA degradation was reported by Van Ginkel and co-workers [145] in semicontinuous activated sludge units which were operated at pH 8^9, whereas at pH 7.0, no EDTA removal was detected even after a long incubation period of 6 months. Hence, EDTA degradation seems to be favoured at moderately alkaline conditions. Also, in sewage treatment plants for wastewaters from pulp and paper industries, an increase of EDTA elimination e¤ciencies from 10 to 50% was observed when the operating pH was raised from the range between 6.5 and 7.0 up to 8.0 or even 9.0 [146,147]. Increasing the pH at which a sewage treatment unit is operated to values between 8.5 and 9.0 to achieve the degradation of EDTA in activated sludge in the absence of support material has now even been patented [148]. Under these conditions, also PDTA was degraded, yet, rather long sludge retention times of at least 2 months were needed for e¤cient PDTA breakdown, while for EDTA removal about 1 week was su¤cient. Eventually, Thomas and co-workers [149] isolated a mixed, EDTA-degrading culture from river water and sludge from an industrial wastewater treatment plant. Among the cultivatable micro-organisms within this culture, representatives of the genera Methylobacterium (35%), Variovorax (17%), Enterobacter (15%), Aureobacterium (11%) and Bacillus (11%) were predominant. However, the ability of the isolated bacteria to grow with EDTA was not tested. Growth of the culture with Fe(III)EDTA was slow, with a doubling time of 66 h, and only 60% of the initial Fe(III)EDTA (5 mM) was degraded. The mixed culture grew also with possible metabolic intermediates such as EDDA, ethylenediamine, NTA and IDA. From this it was concluded that buildup of these compounds was unlikely during growth with EDTA. There is only one study analysing the behaviour of EDTA in a marine ecosystem. In microcosms containing sea water and sediment illuminated with UV light, about 50% of the initially added Fe(III)EDTA complex was converted after 17 weeks [150]. Unfortunately, the poor experimental set-up used does not allow clear distinction between biodegradation and photodecomposition of the Fe(III)EDTA chelate. 3.2.2. In soils and sediments Contradictory results have been published concerning

FEMSRE 706 29-12-00

84

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

EDTA degradation in soils and sediments. Some groups found no biological EDTA breakdown [151,152], whereas others observed a slow microbial EDTA decomposition under aerobic conditions [153^156]. No EDTA mineralisation was found under anaerobic conditions [154,156]. 3.3. DTPA 3.3.1. Wastewater From measurements of EDTA and DTPA concentrations in wastewater e¥uents from pulp and paper mills, in the in£uent and e¥uent of wastewater treatment plants of these mills [157] as well as in the receiving lake [66], it was inferred that DTPA is biologically and/or chemically more readily degradable than EDTA. Indeed, in an activated sludge process, a reduction of DTPA amounting to 50^70% was observed while only 30% of EDTA was eliminated from the wastewater. Also in model wastewater treatment plants, the initial DTPA concentration was reported to be reduced by about 50% [158]. However, elimination could not unequivocally be attributed to microbial activity. In contrast, Hinck and co-workers [140] did not observe DTPA breakdown in activated sludge although the inoculum for the biodegradability tests was taken from a treatment plant that had been in contact with DTPA for over 5 years. 3.3.2. Soils and sediments In soils and sediments, several groups observed the microbial breakdown of DTPA [151,155,156,159] and only Allard and co-workers [152] were unable to detect disappearance of DTPA in sediments. Both NTA and EDTA were formed as products of DTPA breakdown apparently resulting from the cleavage of a C^N bond within one of the ethylenediamine parts of the DTPA molecule [151]. Further metabolites were identi¢ed as incompletely substituted APCAs such as diethylenetriaminetetraacetate, diethylenediaminetriacetate and ED3A which will spontaneously cyclise under neutral and especially acidic conditions to form oxopiperazinepolycarboxylic acids [159]. Unfortunately, it was not possible to distinguish whether these cyclised substances were formed during DTPA transformation itself or only subsequently during the analytical procedure. Nevertheless, the authors proposed that they might be formed during DTPA degradation and then accumulate 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 screening program. The dominant metabolite was ED3A and its corresponding cyclised form [159]. 3.4. Other synthetic APCAs Only one report is available which describes the slow degradation of HEDTA by soil micro-organisms [155] according to which after 173 days of incubation the mea-

sured HEDTA concentration was below 5% of that initially supplied. In sterile controls, the fraction of nonbiologically degraded HEDTA in this experiment was determined as 27%. Nevertheless, recently HEDTA was considered to be biologically not degradable [160]. Typical biodegradation tests using activated sludge as an inoculum and aerobic conditions showed that the amino acid derivatives L-ADA, SDA and MGDA are readily mineralised by micro-organisms. Biodegradation of ASDA was found to be stereospeci¢c and only L-ASDA was easily biodegradable [9]. 3.5. Naturally occurring APCAs 3.5.1. Biodegradation of EDDS In some applications, EDDS is suggested to replace the poorly degradable EDTA. Several biodegradation studies with activated sludge from di¡erent sources have consistently demonstrated the stereospeci¢city of EDDS breakdown [141,161]: whereas [S,S]-EDDS and [R,S]-EDDS disappeared rapidly, [R,R]-EDDS was recalcitrant. However, only the S,S-isomer was completely mineralised [161] and the transformation of [R,S]-EDDS led to the production of a recalcitrant intermediate (AEAA), indicating the removal of one succinyl residue only. This suggests that only [S,S]-EDDS should be employed for the application in both domestic and industrial products. Takahashi and co-workers [141] investigated the degradation of propanediaminedisuccinate (PDDS), a compound very similar to EDDS and also occurring as three di¡erent stereoisomers. In contrast to EDDS, all three isomers of PDDS were easily biodegradable. However, it was not determined whether the compound was completely mineralised. 3.5.2. Other naturally occurring metal-chelating compounds There is little information on the possible biodegradation of naturally occurring APCAs. Nevertheless, several studies indicate that plant-produced chelating organic compounds, such as the APCA mugineic acid, are degraded by rhizosphere micro-organisms under environmental conditions [162^165]. In this way, the microbial £ora may reduce iron uptake of plants by both the degradation of the phytosiderophores and by competition for iron. 4. Degradation of APCAs by pure microbial cultures Up to now, successful isolation of pure, APCA-degrading microbial cultures has been restricted to microbial strains that are able to grow with NTA or EDTA. Some of these isolates have lately been shown to also be able to utilise [S,S]-EDDS (see Section 5.3). Reports on strains utilising DTPA or HEDTA, two other industrially important APCAs, have not been published so far, and we are

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

not aware of any attempts to isolate micro-organisms that grow with representatives of the new generation of APCAs such as L-ADA or MGDA. 4.1. Pure cultures utilising NTA The isolation of NTA-degrading micro-organisms from a wide range of ecosystems has been described including river waters, activated sludge, sediments and soils [166^ 178]. Most of the pure cultures consisted of Gram-negative, obligately aerobic rods capable of utilising NTA as sole source of carbon/energy and nitrogen. Although for the majority of Gram-negative strains detailed characterisation was missing, reports before 1985 mostly allocated them to the genus Pseudomonas [166^170,172^175]. One Gram-negative strain was identi¢ed as Listeria sp. [139]. Two Gram-positive isolates were described and one of them was identi¢ed as a Bacillus sp. [175], whereas for strain TE3 [177] further work is needed to de¢ne its exact taxonomical position (morphological and physiological studies indicate its a¤liation to the Corynebacterium sensu stricto, Mycobacterium, Nocardia, Rhodococcus group [179]). In addition to the strictly aerobic bacteria, denitrifying strains have also been isolated [171,178], whereas ^ so far ^ the enrichment of sulfate-reducing NTA degraders has not been successful [171,180]. A more detailed taxonomical study of Gram-negative NTA-utilising isolates [177^179,181], including most of the isolates still available at the time, revealed that none of them can be assigned to the genus Pseudomonas as de¢ned by de Vos and de Ley [182]. These isolates formed

85

three distinct groups (Table 3): the ¢rst group contains obligately aerobic, motile rods which are mostly pleomorphic and can utilise sugars and also a wide range of other substrates including methylated amines, hence indicating their ability to assimilate C1 units. These strains are phylogenetically localised in the Agrobacterium^Rhizobium branch of the K-subclass of Proteobacteria and they have been combined to form the new genospecies Chelatobacter heintzii [181]. Most recent data indicate that Chelatobacter strains are closely related to bacteria from the genus Aminobacter which are able to grow with methylated amines [183,184]. It is therefore possible that in the future both groups will be combined to a single genus [185]. The second group of isolates (Table 3) consisted of obligately aerobic, non-motile short rods or diplococci unable to grow with sugars. They again could not be assigned to an existing genus and, therefore, were established as the new genospecies Chelatococcus asaccharovorans [181]. This genus belongs to the K-subclass of Proteobacteria and 16S rRNA sequence comparison indicated Rhodopseudomonas as nearest neighbour. The range of APCAs, which can be used as growth substrates by Chelatococcus and Chelatobacter strains, comprised NTA and IDA, while EDTA did not support growth [177]. Moreover, C. asaccharovorans grew with [S,S]-EDDS, whereas Chelatobacter strains did not [186]. No further APCAs have so far been tested as growth substrates for the NTA-degrading strains. The third group of NTA degraders presently consists of 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 Isolation substrate

Strain [reference]

Taxonomical position

Main properties

Respiration

Wmax (h31 ) on isolation substrate

Growth with APCAs other than isolation substrate

NTA

C. asaccharovorans strain TE1 and TE2 [181]

K-2 branch of Proteobacteria

obligately aerobic

0.08^0.1

[S,S]-EDDS

C. heintzii, strains TE4^TE10 and ATCC 29600 and ATCC 27109 [181] strain TE11 [178]

Gram-negative, diplococci, non-motile, with S-layer, not utilising sugars Gram-negative, rods, pleomorphic, motile (1^3 subpolar £agella)

K-2 branch of Proteobacteria, within the Agrobacterium^Rhizobium branch Q-branch of Proteobacteria, Gram-negative, rods, closely related to motile Xanthomonas Gram-negative, rods degrading EDTA only when complexed with Fe(III) K-branch of Proteobacteria Gram-negative, rods, non-motile when grown with EDTA K-branch of Proteobacteria Gram-negative, diplo-rods, within the Agrobacterium^ non-motile Rhizobium branch

obligately aerobic

0.1^0.15

IDA

aerobic and denitrifying

0.08 (aerobic) ; 0.03 (denitrifying)

EDTA

A. radiobacter ATCC 55002 [190]

strain BNC1/DSM 6780 [191] strain DSM 9103 [192]

FEMSRE 706 29-12-00

aerobic

PDTA (only when complexed with Fe(III))

aerobic

0.024

NTA, IDA, ED3A

aerobic

0.06^0.07

NTA, IDA, N,NP-EDDA, [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 inability to grow with IDA, this Gram-negative motile isolate di¡ers only slightly from C. heintzii with respect to its physiological and nutritional characteristics. Polyamine and quinone patterns, however, discriminate TE11 from the latter and suggest an allocation to the Q-subclass of Proteobacteria. The presently available data suggest that TE11 is closely related to but still clearly distinct from members of the genus Xanthomonas [178]. Surface antibodies were raised against C. heintzii and C. asaccharovorans and were used to investigate their distribution in the environment [187,188]. In surface waters, cells reacting with antibodies made up for 0.01^0.1% of the total (acridine orange) cell number. In activated sludges of wastewater treatment plants, the proportion was about 10-fold higher. Surprisingly, no signi¢cant difference in the fraction of cells reacting with antibodies was observed between samples collected from treatment plants, which di¡ered considerably with respect to their e¤ciency in NTA elimination [187]. Despite the immunological detection of bacteria of the genera Chelatobacter and Chelatococcus in activated sludge and in the environment, the contribution of these strains to the degradation of NTA is still unknown. To tackle this question, an experimental approach would be required which consists of the testing of individual cells present in environmental samples for both their serological reaction with the antibodies above described and for NTA degradation. It should be added here that in several Pseudomonas strains (P. aeruginosa, P. £uorescens and P. cepacia), NTA stimulated growth by promotion of the iron uptake, although it did not serve as growth substrate [189]. Apparently, NTA assisted the intracellular incorporation of iron through an active transport system, as do naturally occurring siderophores. In contrast, other chelators such as EDTA, EDDS and EDDHA had an inhibitory rather than a stimulating e¡ect on growth of the tested bacterial strains. Hence, in the environment, NTA might act as a siderophore-like compound for some micro-organisms. However, the interesting question of the intracellular fate of NTA used by micro-organisms as a siderophore has not been further studied. 4.2. Pure cultures utilising EDTA Most likely because of the poor biodegradability of EDTA, the successful isolation of EDTA-degrading bacteria has only recently been reported [190^192] (Table 3). While the isolates described by No«rtemann [191] and Witschel and co-workers [192] exhibit many similarities, the pure culture isolated by Lau¡ and co-workers [190] has rather di¡erent properties with respect the EDTA degradation characteristics. The latter strain, identi¢ed as Agrobacterium radiobacter, was isolated from a secondary waste treatment facility that had received wastes containing Fe(III)EDTA

for several years. Consequently, the isolate grew with Fe(III)EDTA as sole source of carbon, whereas no signi¢cant growth was observed with uncomplexed EDTA [190]. The strain was able to degrade Fe(III)EDTA at initial EDTA concentrations higher than 100 mM but it failed at low concentrations, and the residual EDTA concentrations were always in the range of approximately 3^5 mM. In batch culture, the extent of degradation was more complete for lower initial pHs (probably due to a slower increase of culture pH from excreted NH3 during EDTA breakdown). Addition of peptone or yeast extract to the medium did not a¡ect EDTA degradation but that of glycerol reduced the amount of EDTA degraded drastically. It was suggested that in the presence of glycerol, the Agrobacterium strain utilised EDTA merely as nitrogen source and no longer as carbon and energy source. Other APCAs were tested for growth of the Agrobacterium strain, but only Fe(III)PDTA was degraded and no growth was observed with NTA, IDA or EDDA, although the latter compounds are supposed to be intermediates in the breakdown of EDTA [143,193]. Furthermore, the strain was able to utilise a wide range of sugars, carboxylic acids and amino acids as growth substrates. From an industrial sewage treatment plant receiving EDTA-containing wastewater, No«rtemann [191] was able to isolate a mixed culture of strain BNC1 and BNC2 that grew at considerably lower initial EDTA concentrations (91 mM) than the Agrobacterium strain. In the mixed culture, only strain BNC1 degraded EDTA and its metabolic activity was stimulated in the presence of other micro-organisms probably providing vitamins. Indeed, addition of vitamins was demonstrated to signi¢cantly enhance growth of a pure culture of strain BNC1 [194]. As long as EDTA was complexed with alkaline earth metal ions, strain BNC1 tolerated EDTA concentrations up to approximately 16 mM. However, the presence of uncomplexed EDTA considerably reduced the viability of cells and no growth with free EDTA was observed [194]. A similar observation had already been reported for an NTA-degrading strain [176]. Such adverse e¡ects are probably due to interactions of free EDTA or NTA with metal ions that stabilise the bacterial cell surface because destabilising e¡ects of free EDTA on various Gram-negative bacteria are well known [195,196]. A taxonomic analysis indicated an allocation of strain BNC1 to the K-subgroup of Proteobacteria [193]. However, a more detailed taxonomical study including 16S rRNA sequencing is needed to more exactly allocate this strain, in particular to determine its relation to the other EDTA-degrading strains known so far. Recently, another EDTA-degrading bacterial isolate DSM 9103 has been described [192]. Starting with the mixed culture from Gschwind [144] a Gram-negative, obligately aerobic bacterium was obtained which also belongs to the K-subgroup of Proteobacteria. Both polar lipid pattern and 16S rRNA sequence of the strain are

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

indicative of members of the Agrobacterium^Rhizobium branch. Indeed, the 16S rRNA sequence of the isolate showed highest similarity (96.1%) to that of Rhizobium loti [186], now renamed Mesorhizobium loti [197]. When added to raw wastewater or activated sludge from a municipal wastewater treatment plant cells strain DSM 9103 were able to degrade EDTA to completion (initial concentration approximately 3.5 mM) [192]. However, supplementation with mineral medium (containing high concentrations of alkaline earth metal ions) was necessary to obtain e¤cient EDTA removal in activated sludge, indicating again the importance of EDTA complexation for its biodegradability. Interestingly, strain DSM 9103 is not only able to grow with EDTA but also with other APCAs such as NTA, IDA and [S,S]-EDDS [198]. Although ferrichrome and ferrioxamine type siderophores do not belong to the class of APCAs, it should be mentioned here that pure cultures of ferrichrome Adegrading micro-organisms had been isolated already in the 1960s and that the breakdown of these siderophores has been studied [199^207]. Interestingly, based on 16S rRNA homology, a deferrioxamine B-degrading bacterium was found to be a strain of R. loti [205] and, therefore, it seems a close relative of some of the EDTA-degrading strains presently known (see above). Speciation was found to be a¡ecting degradability with only the iron-free compound (deferrioxamine B, but not ferrioxamine B) being degraded [202,203,206]. 4.3. Potential of bacterial isolates for the treatment of APCA-containing wastewaters NTA is removed e¤ciently during biological wastewater treatment (see Section 3) and, therefore, the application of NTA-degrading isolates for a special treatment of NTAcontaining wastewaters has not been investigated. In contrast, elimination of EDTA in most wastewater treatment plants is negligible [55,56,137]. Thus, the use of EDTAdegrading strains for the treatment of wastewaters containing high concentrations of EDTA was proposed and also patented [208^212]. A process based on a mixed culture of strain BNC1 and BNC2 for treating industrial EDTA-containing wastewaters has been developed. The cells were ¢xed on carrier material and employed in a bio¢lm airlift-loop reactor [211]. The reactor was operated at dilution rates between 0.06 and 0.3 h31 and allowed to achieve an EDTA elimination e¤ciency between 95 and 98% for an EDTA concentration 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 of EDTA and glycerol (each 300 mg l31 ) was fed to the reactor (D = 0.2 h31 ). Initially, when feeding the mixture, no signi¢cant reduction of EDTA elimination e¤ciency was observed but after more than 10 days, the extent of EDTA degradation started to decrease gradually, prob-

87

ably as a result of changes in the composition of the microbial population. In fact, the bio¢lm on the carrier particles was found to become more ¢lamentous and voluminous when feeding the substrate mixture. Hence, in wastewaters containing a high fraction of readily degradable carbon, a two-step process might be of advantage where ¢rst the easily degradable carbon sources are eliminated before EDTA is removed with the help of strain BNC1 in a second step [212]. Industrial wastewaters often contain the Fe(III) complex or other heavy metal complexes of EDTA, but strain BNC1 is not able to attack these chelates (see Section 7). To treat wastewaters containing such EDTA complexes, a precipitation step was established before the EDTA solution was supplied to the bioreactor. By the addition of a large excess of calcium hydroxide to a Fe(III)EDTA-containing synthetic wastewater, Fe3‡ was exchanged against Ca2‡ and then precipitated as iron hydroxide. A solution containing mainly CaEDTA was obtained, and after adjusting the pH, the Fe(III)EDTA elimination e¤ciency by strain BNC1 was about 98% [213]. Eventually, the system was tested for its performance in eliminating EDTA in real wastewaters originating from the dairy industry, vat cleaning in a brewery, the pulp and paper industry, and from the photographic industry. The data indicated that all these wastewaters were treatable with immobilised EDTA-degrading bacteria but occasionally speci¢c pre-treatment steps were required. EDTA in wastewater from pulp and paper manufacturing was e¤ciently removed without any pre-treatment. Wastewaters from the dairy industry could be treated in a two-step process because of the presence of high amounts of non-EDTA DOC, which must be removed ¢rst. Wastewaters from vat cleaning or the photographic industry again contain high amounts of heavy metals and a precipitation step (or a similar process to remove these metals) has to be carried out before biological treatment is possible [212]. The results are quite promising with respect to a potential utilisation of immobilised cells of strain BNC1 for the biological treatment of EDTA-containing wastewaters. Nonetheless, a note of caution should be added because all tests for EDTA degradation in industrial wastewaters have so far been performed under constant conditions, a situation that is rarely encountered in real life. Therefore, important data are still missing with respect to overall elimination e¤ciencies in case of variations in wastewater composition or other £uctuations of parameters (e.g. in pH or feeding rate). 5. Biochemistry and genetics of APCA degradation Information on the biochemistry and genetics of APCA degradation is of course restricted to those compounds for which isolation and cultivation of pure bacterial cultures has been reported, i.e. NTA, EDTA and [S,S]-EDDS.

FEMSRE 706 29-12-00

88

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

Again, of all APCAs, NTA has clearly received most attention. 5.1. NTA degradation Research mainly focused on the intracellular metabolism of NTA and the question of its transport into cells was rarely addressed. One can merely speculate that NTA is actively transported across the cell membrane because [14 C]NTA uptake was inhibited by KCN [169,214]. 5.1.1. NTA catabolism Based on identi¢cation of products accumulated during

NTA degradation and of enzymatic activities enhanced in cell-free extracts of NTA-grown cells, already some 25 years ago a pathway for the aerobic NTA breakdown was proposed (Fig. 4) [166,170,174,215] : NTA was thought to be metabolised in two steps via IDA to glycine and glyoxylate, with the ¢rst step being catalysed by a monooxygenase (MO) [166,170]. The products were then channeled into the central metabolism. A comparison of enzyme activities in cell-free extracts of glucose- and NTAgrown cells gave strong evidence for the involvement of glycine decarboxylase and serine hydroxymethyl transferase in the subsequent metabolism of glycine, and that glyoxylate 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) hydroxypyruvate reductase ; (14) glycerate kinase ; (methylene)FH4 , (N5 ,N10 -methylene) tetrahydrofolic acid; PLP, pyridoxal phosphate ; TPP, thiamine pyrophosphate (from [6] with permission).

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

nase (DH) took care of the transformation of glyoxylate into glycerate [170]. The proposed pathway was con¢rmed by puri¢cation and characterisation of two enzymes from C. heintzii ATCC 29600 that catalyse the degradation of NTA and IDA, respectively [216,217]. A MO (NTA MO) was found to catalyse the attack on NTA leading to the formation of IDA and glyoxylate. In a second step, IDA is oxidatively split into glycine and glyoxylate, not by NTA MO as originally supposed [166], but by a membrane-bound IDA DH. Both enzymatic activities were not only detected in C. heintzii but also in Chelatococcus strains [218].

89

from NADH2 to FMN (Fig. 5). cB can be replaced by other NADH2 :FMN oxidoreductases, for instance the oxidoreductase from Photobacterium ¢scheri which normally supplies FMNH2 to the luciferase enzyme. Therefore, FMN and FMNH2 most presumably act here as real coenzymes not being tightly bound to cB. Recently, similar two-enzyme systems have been described, also consisting of a NADH2 :FMN oxidoreductase and a MO, namely those involved in the oxidation of dibenzothiophene [221] and in the synthesis of antibiotics such as pristinamycin IIA [222], actinorhodin [223] or valanimycin [224]. It therefore is not surprising that cA and cB exhibited the highest homology with proteins of this two-enzyme system group when the amino acid sequences were compared [220, 224,225]. Eventually, sequence comparison revealed a signi¢cant similarity between FMNH2 -dependent MOs from E. coli, Bacillus subtilis and Pseudomonas putida catalysing the desulfonation of alkanesulfonates and several of the above mentioned MOs (including NTA MO) [226]. In this study, also a conserved domain present in all MOs was detected which might be part of the active site of these enzymes. This domain, however, contained none of the protein motifs known so far. Also from Chelatococcus strain TE2, a FMNH2 -dependent MO was puri¢ed which exhibited many similarities with cA from C. heintzii ATCC 29600. The N-termini of

5.1.1.1. Characteristics of NTA MO. NTA MO from C. heintzii ATCC 29600 was ¢rst puri¢ed and characterised by Uetz and co-workers [216] and later more information was obtained by sequencing the corresponding genes [219,220] (Table 4). NTA MO was originally thought to consist of two weakly associated components, cA and cB, both of which must be present to catalyse the transformation of NTA [216]. More recently, it was demonstrated [220] that cA and cB are two distinct enzymes, with cA being the true MO catalysing the oxidative cleavage of NTA with the consumption of FMNH2 and molecular oxygen, and cB being an oxidoreductase providing FMNH2 for the MO by transferring reducing equivalents Table 4 Some properties of key enzymes involved in APCA degradation NTA MO (cA) from C. heintzii

NTA DH from strain TE11

IDA DH from C. heintzii

EDTA MO (cAP) from strain DSM 9103

EDDS lyase from strain DSM 9103

Location Molecular mass SDS (kDa) Native (kDa) Substrate

Soluble

Soluble

Membrane-bound

Soluble

Soluble

47 99 NTA

80 170 NTA

^ ^ IDA

44 210 EDTA

Products

IDA+glyoxylate

IDA+glyoxylate

Glycine+glyoxylate

Further substrates

^

^

^

Cofactors required Redox component Electron acceptor

FMNH2 ^ O2

^ (cyt b)b Respiration chain, ubiquinone pool

Accessory enzymes

NADH2 :FMN oxidoreductase (cB) MgNTA (CoNTA ZnNTA, MnNTA, NiNTA, Fe(III)NTA)

PMSa FAD PMS, in vivo acceptor unknown NtrR (under anoxic conditions) Uncomplexed NTA

ED3A, N,NPEDDA+glyoxylate DTPA, HEDTA, NTA, PDTA FMNH2 ^ O2

58 130 [S,S]-EDDS, [R,S]-EDDS AEAA+fumarate

Required speciation of substrate

Km (substrate)

0.5 mM (for MgNTA)

0.095 mM (for NTA DH/NtrR complex) ; 0.19 mM (for NTA DH only, under aerobic conditions)

Uncomplexed IDA (enzymatic degradation also in presence of Mg, Ca, Ba and Mn) 8 mM (in presence of Ca)

Data were taken from [198,216,217,220,229,235,251,252]. a In vivo cofactor unknown. b Indicated by spectral properties.

FEMSRE 706 29-12-00

NADH2 :FMN oxidoreductase (cBP) MgEDTA (ZnEDTA, MnEDTA, CoEDTA)

^ ^ ^ ^

Uncomplexed EDDS (MgEDDS, BaEDDS, CaEDDS, MnEDDS)

0.8 mM (for MgEDTA) 0.2 mM

90

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

both proteins were virtually identical with only one amino acid being di¡erent. In addition, the overall amino acid composition of the two MOs was very similar (however, the whole amino acid sequence of cA from C. asaccharovorans has not been unraveled yet). In contrast, no protein equivalent to cB has yet been isolated from C. asaccharovorans, but the existence of such a protein is compulsory as indicated by the fact that cB from C. heintzii could form a functional, NTA-oxidising enzyme complex with cA from C. asaccharovorans [218]. Consistently, studies with antisera raised against cA and cB from C. heintzii ATCC 29600 indicated the presence of cA in both Chelatobacter and Chelatococcus strains, while cB was found to be restricted to Chelatobacter strains [6,218]. cA and cB could immunologically only be detected in cells which were grown with NTA. Interestingly, in trimethylamine-cultivated C. heintzii bacteria, a protein of 70 kDa, thus signi¢cantly larger than NTA MO, crossreacting with cA antiserum was found. In cell extracts, however, no NTA MO activity but only trimethylaminestimulated NADH2 oxidation was detectable [218]. Also DNA^DNA hybridisation data suggested that the gene encoding cA is more conserved than the cB gene [225] (see Section 5.1.4). This con¢rms the crucial role of cA (NTA MO) for NTA degradation in contrast to that of cB, which probably can be replaced by any other enzyme, providing reduced FMN. Here it is noteworthy that ^ apart from the luciferase system and the two-enzyme systems listed above ^ other NADH2 :FMN oxidoreductases providing free reduced £avins have been described which seem to be involved in the reduction of iron complexed to siderophores [227]. However, no similarity at the amino acid sequence level among the various classes of these oxidoreductases has so far been found. An unusual property of the oxidoreductase cB from C. heintzii was that the reduction of FMN was highly stimulated by the addition of NTA. Other structurally related compounds, among them also IDA, had no such stimulating e¡ect when tested in reaction mixtures containing both cB and cA [216]. According to these NADH2 oxidation assays, the substrate spectrum of the NTA MO (cA) was stated to be rather narrow. Yet, catalysis by the NTA MO of the oxidative splitting of some of those compounds that did not stimulate the NADH2 oxidation by cB cannot be totally excluded. Nevertheless, IDA is certainly not further attacked by the NTA-degrading enzyme system as originally proposed by Firestone and co-workers [164] since no glyoxylate is formed from IDA in the presence of cA and cB [216]. 5.1.2. Enzymatic degradation of IDA Since the NTA MO does not catalyse the cleavage of IDA, an additional enzyme must exist. Indeed, IDA-oxidising activity was detected in the particulate membrane fraction from cells of C. heintzii and C. asaccharovorans, which was subsequently attributed to a membrane-bound

IDA DH [217] (Table 4). IDA DH from C. heintzii was distinctly di¡erent from other membrane-bound DHs such as succinate DH and it merely catalysed the oxidative splitting of IDA to glycine and glyoxylate. IDA DH is probably not only associated with the membrane but exists as an integral membrane protein which feeds electrons from the oxidation of IDA into the electron transport chain via the ubiquinone pool [217]. Partial enrichment of IDA DH was achieved by extracting the enzyme from membranes with the help of cholate and incorporating it into soybean phospholipid vesicles. In this arti¢cal system, IDA DH activity was fully reconstituted upon addition of the major quinones in the Chelatobacter genus as intermediate electron carriers, i.e. ubiquinone Q1 or ubiquinone Q10 , and of iodonitrotetrazolium chloride as terminal electron acceptor [217]. 5.1.3. NTA degradation in denitrifying bacteria Obviously, the initial step in the metabolism of NTA in denitrifying NTA degraders has to proceed via an oxygenindependent step. In membrane-free extracts of strain TE11 grown under anoxic conditions, a protein complex consisting of two enzymes, a NTA DH and a nitrate reductase (NtrR), catalysed the ¢rst step of NTA oxidation resulting in the formation of IDA and glyoxylate [228] (Table 4). Only with both enzymes present, NTA was transformed, and activity was coupled to a phenazinemethosulfate (PMS)-dependent transfer of electrons from NTA to nitrate which was reduced to nitrite. Apart from the arti¢cial dye PMS, no naturally occurring compound has been found so far which mediates the electron transfer from NTA DH to NtrR. As for NTA MO, the substrate spectrum of the DH seems to be restricted to NTA. N-Ethylmaleimide, a thiol-binding reagent, inhibited NTA DH but an excess of dithiothreitol partly restored enzymatic activity, indicating the involvement of thiol groups in reaction catalysis by NTA DH [229]. As a redox component NTA DH contains a covalently bound FAD moiety. Additionally, an iron content of about four iron atoms per mol of enzyme was found. Since results from di¡erence spectra of oxidised versus reduced enzyme argued against the presence of heme chromatophores, NTA DH is supposed to contain two 2Fe^2S clusters, one in each monomer of the homodimeric enzyme [229]. Thus, electrons derived from NTA are presumably ¢rst transferred to FAD, then to the iron^sulfur cluster, and ¢nally to PMS or the so far unknown in vivo electron carrier. The in vitro found NTA DH/NtrR complex is rather unusual and in fact the ¢rst enzyme complex reported where a catabolic DH is so tightly associated with a NtrR. Therefore, the in vivo localisation of both NTA DH and NtrR was investigated by immunochemical labelling of the enzymes to address the question of whether the complex is an artefact produced during cell disruption or whether it really exists in the cells. NTA DH was detected in the cytoplasm, whereas NtrR was associated with or

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

integrated in the cytoplasm membrane [225]. The reductase from the NTA DH/NtrR complex appears to belong to the dissimilatory type of NtrRs [225] because NtrR expression was not observed under conditions where only the assimilatory type of NtrRs should be produced, i.e. during aerobic growth with nitrate as sole nitrogen source. In contrast, ammonium did not repress the reductase expression under anoxic conditions. Moreover, NtrR production was independent of the presence of NTA in the growth medium while NTA DH was only formed during cultivation with NTA [225]. NTA DH was also puri¢ed from aerobically grown strain TE11, and seemed to be identical to the enzyme expressed under anoxic conditions [229]. However, during aerobic growthm, electrons were transferred to the respiration chain, but again, only in the presence of PMS. Although the NTA DH/NtrR complex from denitrifying cells and NTA DH from aerobically grown cells catalysed the same overall reaction, association of NTA DH with NtrR led to a marked change in the kinetic properties with an increase in a¤nity for NTA [6]. However, it remains to be tested whether this is real or an artefact originating from the puri¢cation procedure. The interesting aspect of the process of in vivo electron transfer from NTA DH to either NtrR or to the respiration chain is a still unsolved and intriguing question, especially when bearing in mind that NTA DH apparently is not associated with the cytoplasmic membrane. In analogy to the situation found in case of the functionally related trimethylamine DH [230], an electron-transferring £avoprotein was suggested to be involved [229], but this hypothesis has not yet been con¢rmed. Also in the denitrifying strain TE11 ^ similar to the obligately aerobic Chelatobacter and Chelatococcus strains ^ IDA is further oxidised by a membrane-bound IDA DH [229]. 5.1.4. Genetic information concerning NTA metabolism Quite recently, the NTA-metabolising enzymes described above were also investigated at the genetic level [219,220,225]. The genes of the two enzymes, cA and cB (ntaA/nmoA and ntaB/nmoB, respectively), were cloned from C. heintzii ATCC 29600. They were oriented divergently with an intergenic region of 307 bp, a rather unusual organisation for genes whose products act so closely together. Downstream of the gene for cA, additional open reading frames (ORFs) (ORF1/nmoR and nmoT) were found. In its N-terminus, the 24.4 kDa gene product of ORF1/nmoR exhibited a DNA-binding motif (helix-turnhelix) which is characteristic for a family of bacterial regulatory proteins called GntR family. Hence, the ORF1/ nmoR gene product might be involved in the regulation of the expression of the genes for cA and cB. Unfortunately, the attempts to further characterise the putative NTA regulator have failed so far. Sequence similarities

91

of nmoT with several transposases indicate that it may be part of an insertion element. To test the presence of genes homologous to ntaA and ntaB in other NTA-degrading strains, DNA^DNA hybridisation experiments were performed using di¡erent restriction fragments from cloned parts of the cA and cB genes from C. heintzii ATCC 29600 as probes [225]. Surprisingly, in C. heintzii TE6, only regions homologous to the cA gene and ORF1 but not to the gene for cB were revealed although previous immunological studies had indicated the presence of proteins similar to cA and cB [218]. Conversely, no crossreacting proteins similar to cB were detected in C. asaccharovorans TE2, whereas at the DNA level, regions highly homologous to those of ORF1 and the genes encoding cA and cB in C. heintzii ATCC 29600 were found. However, shortly after the start of the ntaB equivalent gene from strain TE2, the homology with ntaB from C. heintzii ATCC 29600 ended. Apparently, the gene for cB in C. asaccharovorans TE2 is disrupted in its reading frame, and most probably no intact cB is synthesised. An attempt was made to clone the gene for NTA DH from denitrifying strain TE11 [225] starting with oligonucleotide probes derived from the N-terminal amino acid sequence of the enzyme [229]. However, because the speci¢city of the available sequence information was too low, the probes synthesised hybridised with many di¡erent DNA fragments. This is consistent with the observation that the N-terminal amino acid sequence of NTA DH exhibited high similarity to other £avin-containing enzymes all possessing a common FAD-binding segment (LKL-fold) in their N-terminus [229]. 5.2. Catabolism of EDTA and the similarity of EDTA MO with NTA MO Due to the successful isolation of EDTA-degrading bacterial strains within the 1990s, ¢rst information on the metabolism of EDTA could be obtained. However, so far only strain BNC1 and DSM 9103 have been studied biochemically while the pathway for EDTA breakdown in the Agrobacterium strain growing with Fe(III)EDTA is still unknown. Uptake of EDTA into cells of strain DSM 9103 was mediated by an energy-dependent carrier, which is most probably driven by the proton motive force. This transport system had a high apparent a¤nity for EDTA. It was rather speci¢c for EDTA because, of several APCAs and other structurally related compounds, only DTPA competitively inhibited the transport of EDTA [231]. Based on products excreted by a mixed microbial culture during EDTA degradation, Belly and co-workers [143] proposed two possible pathways for the microbial breakdown of EDTA, i.e. either the successive removal of acetyl groups from EDTA or the cleavage of a C^N bond within the ethylenediamine part of the molecule. The former mechanism, the removal of acetyl groups, was re-

FEMSRE 706 29-12-00

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 BNC1 [193,232] and from strain DSM 9103 [198]. The EDTA-splitting activity required molecular oxygen, NADH2 and FMN and yielded glyoxylate and ^ transiently ^ ED3A, the latter being further degraded under the same assay conditions resulting in the formation of another glyoxylate and N,NP-EDDA. In both strains, no N,NEDDA production was observed, indicating regiospeci¢c removal of the second acetyl group from EDTA. From strain DSM 9103, a two-enzyme system was puri¢ed which catalysed the removal of two acetyl groups from EDTA (Fig. 5, Table 4) [198]. Most likely, one enzyme (cAP) is a MO catalysing the oxidative cleavage of EDTA and ED3A while consuming oxygen and reduced FMN (it has not been demonstrated yet that molecular oxygen is incorporated into water and glyoxylate). The second enzyme (cBP) is a NADH2 :FMN oxidoreductase that provides FMNH2 for the MO. As in case of NTA MO, the oxidoreductase from P. ¢scheri could take over the function of cBP. Furthermore, cB from the NTA-oxidising enzyme system was able to replace cBP in the EDTA-degrading enzyme system, and vice versa. In contrast to the NTA-degrading enzyme system, EDTA MO exhibited a broader substrate spectrum. It cleaved not only EDTA and ED3A but also other APCAs such as NTA, DTPA, HEDTA, PDTA and 1,3-diamino-

propanoltetraacetic acid. N,NP-EDDA was not further transformed by EDTA MO but a cofactor-independent N,NP-EDDA-degrading activity was detected in cell-free extracts of strain DSM 9103 [198]. Additionally, N,NPEDDA transformation was also observed in membrane fractions of DSM 9103 cells. By both, the soluble and the particulate fraction, N,NP-EDDA was oxidatively split resulting in the formation of one glyoxylate molecule. Apparently, an additional acetyl residue was removed from the ethylenediamine backbone probably leaving EDMA (Fig. 5). Preliminary data indicate the involvement of a DH [233]. However, the enzyme(s) responsible for this activity still remain(s) to be isolated. One can speculate that EDMA is further metabolised by elimination of the last acetyl moiety leaving ethylenediamine which can be transformed to glycine. Alternatively, EDMA might be cleaved within the ethylenediamine part of the molecule yielding ¢nally two molecules of glycine. As a third possibility, EDMA might be converted into IDA by the removal of the primary amine group. IDA can then be oxidised to form glycine and glyoxylate. The metabolic pathway for the products of EDTA degradation, glycine and glyoxylate, is assumed to be similar to that described previously for the NTA degrader C. heintzii [233]. All in all, important similarities between the NTA me-

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

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 acting together with a distinct oxidoreductase providing reduced FMN. 5.3. EDDS catabolism EDDS is a structural isomer of EDTA and it can also be used as a growth substrate by the EDTA-degrading strain DSM 9103. Therefore, it was speculated that EDDS might be transported into the cells by the EDTA carrier and subsequently be cleaved by EDTA MO. This, however, proved not to be the case [198,231,235]. Whereas uptake of EDDS by strain DSM 9103 has not yet been studied, the initial step in the intracellular EDDS breakdown has been elucidated [235]. The catabolism of EDDS was initiated by a carbon^ nitrogen lyase catalysing the non-hydrolytic cleavage of the C^N bond between the ethylenediamine part of the molecule and one of the succinyl residues without any cofactors being required (Fig. 6, Table 4). The reaction led to the formation of fumarate and AEAA. Also in C. asaccharovorans, a similar [S,S]-EDDS-splitting activity was detected requiring no cofactors and resulting in the formation of the same products as those found in DSM 9103, thus indicating the action of the same type of enzyme [235]. The further degradation of AEAA remains still to be unraveled. To date, one can merely speculate that, catalysed by a DH or a MO, the C^N bond between the succinyl residue and the ethylenediamine part of the molecule is split, or that an aspartyl residue is removed by the cleavage of a C^N bond within the ethylenediamine part of AEAA [233]. Out of the three stereoisomers of EDDS ([S,S]-, [R,R]and [R,S]-EDDS), the lyase accepted only [S,S]- and [R,S]EDDS. Probably only the S-con¢guration of the chiral Catom can be attacked by the enzyme. Apparently, the

93

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 the true substrate for the lyase [235].

enzymatic degradation of AEAA is also stereospeci¢c [161]. While a microbial consortium mineralised [S,S]EDDS totally, utilisation of [R,S]-EDDS resulted in the formation of a dead-end product, identi¢ed as AEAA which was supposedly present in the D-con¢guration. Consequently, only l-AEAA, which should be formed during [S,S]-EDDS breakdown, can be expected to be further metabolised. 6. Regulation of the APCA degradation and enzyme expression 6.1. Regulation of NTA degradation The ability to metabolise NTA was inducible in Chelatobacter, Chelatococcus and denitrifying strains. NTA- and IDA-degrading activity was only detected in cell-free extracts and membrane fractions, respectively, of bacteria grown with either NTA or IDA [100,209]. This, however, provides only limited information concerning the regulation processes taking place under conditions found in wastewater treatment plants or in the environment which are characterised by rather low substrate concentrations. Therefore, experiments with C. heintzii ATCC 29600 were conducted in carbon-limited continuous culture fed with either NTA or glucose to get a better insight in possible regulation mechanisms of NTA breakdown during growth under conditions similar to those found in the environment [236,237]. In a glucose-limited continuous culture, i.e. in the absence of NTA, both the NTA- and IDA-oxidising activ-

FEMSRE 706 29-12-00

94

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

ities were very low and close to the detection limit, independent of the growth rate applied. When cultivated with NTA as sole source of carbon and nitrogen, the speci¢c activity of NTA MO increased with decreasing dilution rates. The cells exhibited a maximum speci¢c activity (approximately 180 Wmol NTA (g protein)31 min31 ) at dilution rates between 0.02 and 0.03 h31 , whereas at a dilution rate close to Wmax of C. heintzii, the speci¢c NTA MO activity was 4^5 times lower. Thus, at high growth rates with NTA, repression of the NTA MO activity was observed, a pattern often found for enzymes catalysing the ¢rst step in catabolic pathways [238,239]. Under the same conditions, IDA DH activity was approximately constant at growth rates above 0.03 h31 and only decreased considerably at lower growth rates. In both nature and wastewater treatment plants, NTAdegrading micro-organisms will not only encounter NTA as sole substrate but they will grow in the presence of a complex mixture of potential carbon and/or nitrogen sources [240]. Indeed, under laboratory batch conditions, both Chelatobacter and Chelatococcus spp. are capable of consuming NTA in combination with a suitable carbon substrate generally resulting in an enhanced speci¢c growth rate [177]. To investigate the degradation of NTA during mixed substrate growth in more detail, C. heintzii was cultivated with di¡erent mixtures of glucose and NTA in a carbon-limited chemostat and special attention was given to the expression of enzymes involved in NTA metabolism [236]. Synthesis and activity of NTA-degrading enzymes was controlled by the ratio of substrates rather than their actual concentrations in the feed. During growth with mixtures containing less than 1% of the total carbon as NTA, the amount of cA and cB produced was close to the detection limit. Although the speci¢c activities of NTA MO and IDA DH in cell extracts were very low, the bacteria were able to degrade the small amounts of NTA provided. Either the constitutive enzymatic activity was su¤cient to consume the NTA supplied or an alternative, currently unknown enzyme system was involved. Feeding a NTA/glucose mixture containing 3.6% of NTA^carbon triggered induction of NTA enzymes above the constitutive background level and a further increase of the NTA fraction resulted in clearly enhanced activities of both NTA MO and IDA DH. Assuming that under environmental conditions NTA-degrading enzymes are similarly regulated by the proportion of NTA to the total carbon utilised, no signi¢cant induction of these enzymes can be expected in rivers receiving a high load of treated wastewater where NTA may contribute only to some 0.1^1%. In wastewater treatment plants, however, NTA might contribute to some 1^10% of the total utilisable carbon, and here, induction of NTA-degrading enzymes can be expected to play a signi¢cant role. Obviously in the engineered or natural environment, NTA-degrading bacteria will rarely experience steady-state conditions, but rather changes in substrate concentration

and spectrum caused by shock loading, diurnal and weekly cycles, heavy rainfalls, etc. Consequently, cells have to adapt to the changing conditions, and the lag phase resulting from these adaptation processes can severely a¡ect the e¤ciency of biodegradation over long time periods [241]. To investigate the dynamic behaviour of NTA degradation in a continuous culture of C. heintzii under transient conditions, the feed was switched from a medium containing glucose only to one containing either NTA only or mixtures of glucose and NTA [237]. When glucose-pregrown cells were suddenly confronted with NTA as sole source of carbon, nitrogen and energy, a long lag phase of about 25 h was measured until NTA MO expression started. This lag phase was considerably shortened when the cells were supplied with mixtures of NTA and glucose instead of NTA only. Possibly in the latter case, the portion of glucose available after the switch was able to support the rearrangement of the cellular metabolism, particularly to provide energy and building blocks for the synthesis of NTA-degrading enzymes. Hence one can expect that alternative carbon substrates have a considerable positive e¡ect on the expression of other enzymes under changing environmental conditions. Although such steady-state and dynamic experiments with pure cultures of C. heintzii provide some basic information on the behaviour of this bacterium, one still cannot infer how NTA degradation is really regulated in wastewater treatment plants or in surface waters. In particular, the question of whether regulation proceeds at the level of enrichment of NTA-degrading micro-organisms or rather at the level of induction/repression of the pollutantspeci¢c enzyme systems remains unanswered. Nevertheless, a number of observations suggest that for shortterm £uctuations in the supply with NTA regulation takes place at the enzyme level [188]. Firstly, when NTA was added to activated sludge in laboratory batch cultures, only expression of NTA MO but no enrichment of NTA-degrading bacteria from the genera Chelatobacter and Chelatococcus (assessed with surface antibodies) was detected within about 100 h. Also, only a slight rise of Chelatobacter bacteria was observed after activated sludge in porous pot plants that had been fed with synthetic wastewater containing high concentrations of NTA for more than 1 month. Secondly, no enrichment of C. heintzii was found in the Swiss wastewater treatment plant on the mountain Saentis which was periodically exposed to unusually high in£owing NTA concentrations. NTA MO, however, became detectable within a few days after the ratio of NTA^carbon to DOC had increased over 10% in the in£uent, resulting in a markedly higher NTA removal e¤ciency [188]. In contrast to these data [188], in situ hybridisation with 16S rRNA probes speci¢c for bacteria from the Rhizobiaceae group revealed a steady raise of Rhizobiaceae in a fed-batch culture of activated sludge daily amended with 1 g l31 NTA [242]. Most of the APCA-degrading strains

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

known to date belong to this family of Rhizobiaceae. However, cells detected by the Rhizobiaceae probe did not react with probes speci¢c for C. heintzii, which might indicate that important NTA-degrading bacteria from the Rhizobiaceae group are still unknown. Unfortunately, in these studies, measurements of NTA concentrations or enzyme activities in the culture were lacking to prove whether or not the observed enrichment was really linked to the degradation of NTA. 6.2. Regulation of degradation of EDTA and EDDS As for the NTA-degrading enzymes, EDTA MO is also inducible in the two EDTA-degrading strains BNC1 and DSM 9103. When strain DSM 9103 had been grown with EDTA or ED3A, the rate of EDTA consumption in cellfree extract was stimulated 5^10-fold over the constitutive level of bacteria cultivated with glycerol. Growth with NTA or IDA, however, did not result in such a stimulation [198]. This partly contrasts with the observations reported for EDTA-degrading strain DSM 9103: compared to EDTA-grown cells, both EDTA transport and EDTAoxidising activity were 10 times lower when cells had been grown with IDA or fumarate, whereas it was similarly high (about 80% of the activity of EDTA-grown cells) when the strain had been cultivated with NTA or [S,S]EDDS. This also suggests that transport and catabolic enzymes are regulated in a co-ordinate manner [198,235]. Likewise, [S,S]-EDDS-degrading activity in the two strains DSM 9103 and C. asaccharovorans was inducible. Lyase-catalysed EDDS transformation was only detected in cell extracts of [S,S]-EDDS-grown bacteria but not in fumarate-, EDTA- or NTA-grown cells [235]. 7. In£uence of metal speciation on microbial APCA degradation In the previous chapters on microbial and enzymatic APCA degradation, one important point was not touched, namely the fact that in both growth media and the environment APCAs are generally complexed with metal ions. Consequently, metal ions must be expected to have a substantial in£uence on the biodegradability of APCAs. In this context, several aspects have to be considered: initially, the APCA molecule has to be transported across the cytoplasm membrane metal-free or in association with a metal ion. If the latter were the case, a cell would encounter an enormous in£ux of metal ions. To deal with this situation, the cell can either excrete the metal ions or inactivate the cations by precipitating them intracellularly with a suitable anion (e.g. phosphate). On the other hand, if only the free APCA molecule enters the cell, then the metal has to dissociate from the complexing agent prior to uptake. This could be achieved by destabilising the metal^ APCA complex during transport, resulting in the release

95

of the metal ion at the outer side of the cytoplasmic membrane. Alternatively, the cells could transport only free APCA existing in equilibrium with the metal-complexed APCA in the surrounding medium. This question of APCA speciation arises also at the level of enzyme-catalysed breakdown in the cytoplasm. As to the properties of the metal^APCA species itself, one is left with the question of whether the thermodynamic stability of a complex, its dissociation kinetics, or its structure governs the uptake into a cell or the enzymatic transformation. Several di¤culties are encountered when investigating the e¡ect of APCA speciation, in particular in assays with whole cells. Firstly, metal toxicity has to be considered. Secondly, the presence of metal ions as well as metalcomplexing ligands on the bacterial surface should be taken into account, because both might signi¢cantly a¡ect the complexation of the APCA added to the system. However, it is di¤cult to assess the in£uence of these surface-bound ligands or metals. Thirdly, although the speciation of APCAs at the beginning of an assay can be predicted rather easily, changes in the course of such an experiment are di¤cult to assess. 7.1. In whole cells 7.1.1. NTA Investigations of the biodegradability of di¡erent NTA species were ¢rst performed in the 1970s. The systems studied (activated sludge, river water, soil) were rather complex ^ with many other ligands and metal ions being present beside the NTA complex ^ and, therefore, no conclusive information was obtained. Nevertheless, the data suggested that NTA complexed with Ca2‡ , Fe3‡ , Mn2‡ or Pb2‡ is probably easily biodegradable under environmental conditions whilst NTA bound to Cd2‡ , Cu2‡ , Hg2‡ or Ni2‡ might be more recalcitrant [132,243^246]. Somewhat later, degradation of metal^NTA complexes by a `Pseudomonas' species (now C. heintzii ATCC 29600) was studied under more de¢ned conditions [247]. However, also here bu¡er systems were used which themselves have complexing properties and, therefore, the speciation of NTA was not precisely de¢ned. When metal concentrations were kept low enough to avoid toxicity, the bacterial strain was able to attack Ca2‡ , Mn2‡ , Mg2‡ , Cu2‡ , Zn2‡ , Cd2‡ and Fe3‡ chelates of NTA, as well as free NTA as determined by oxygen consumption and 14 CO2 evolution from 14 C-labelled NTA. Only NiNTA was resistant towards degradation. Recently, more elaborate data on the e¡ects of NTA speciation on microbial degradation were presented in which calculations of the chemical speciation were included [248]. Degradation assays with C. heintzii ATCC 29600 revealed ¢rst order kinetics for the decomposition of several degradable NTA species at concentrations ranging from 0.05 to 5.23 WM NTA with degradation rates in the

FEMSRE 706 29-12-00

96

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

following order: H2 NTA3 s CoNTA3 VFeOHNTA3 V ZnNTA3 s AlOHNTA3 s CuNTA3 s NiNTA3 . This order indicates no relationship with the chemical stability of the di¡erent complexes. However, a relationship between the rates of degradation and the lability of the various metal^NTA complexes was established. This implies that the rates of formation of H2 NTA3 from MeNTA, i.e. the dissociation rates of the complexes, determine the degradation rates. Two models can explain these results: the ¢rst model is based on the assumption that only H2 NTA3 is taken up into the cells, with the extracellular dissociation of the metal^NTA complex being the rate-limiting step in the degradation. The second model assumes that all species, i.e. H2 NTA3 and metal^NTA complexes, enter the cytoplasm, where the species accepted by the NTA MO has then to be generated. Previously, Uetz and coworkers [216] had demonstrated that the preferred substrate of NTA MO is MgNTA, whereas CaNTA or H2 NTA3 are not accepted by the enzyme. Therefore, an exchange of Mg2‡ for the metal complexed to NTA must take place intracellularly before NTA degradation can proceed. This exchange reaction will be the rate-limiting step which is again dependent on the dissociation kinetics of the metal complexes. Unfortunately, the data collected are not su¤cient to establish which of the processes actually occurs. For this, short-term studies of NTA uptake into cells, preferably carried out with membrane vesicles, are required. Such data are still lacking and experience from our laboratory indicates that it will be very di¤cult to prepare membrane vesicles from C. heintzii [214]. Somewhat contrasting results were reported by Madsen and Alexander [139]. Based on speciation models, de¢ned media were prepared whose composition favoured the presence of one predominant species of NTA. Under such conditions, an unde¢ned microbial consortium mineralised CaNTA but not AlNTA, MgNTA, Fe(III)NTA or free NTA. The same pattern was found for a Listeria sp. isolated in the course of the study although the isolation medium contained mainly free NTA. Con¢rmatory tests with this strain in the presence of constant NTA and varying Ca concentrations (9 NTA concentration) showed that the extent of NTA mineralisation increased with increasing Ca concentrations. 7.1.2. EDTA Several reports on the degradability of various EDTA complexes su¡er from the same drawbacks as those described above for NTA, i.e. that unde¢ned, rather complex systems were used. In such systems, EDTA was found to be degraded by mixed microbial consortia when added in the free form or as chelate of Cu2‡ , Cd2‡ , Zn2‡ , Ni2‡ , Mn2‡ , Ca2‡ or Fe3‡ [149,153]. Contrasting requirements of pure EDTA-degrading cultures with respect to EDTA complexes supporting growth were reported. On the one hand, growth of an Agrobacterium sp. was only observed when EDTA was supplied as

Fe(III)EDTA. Free EDTA as well as complexes with Ni, Cu or Co(III) did not support growth [190,249]. On the other hand, the EDTA-degrading strain BNC1 was not capable of utilising Fe(III)EDTA [194]. Growth of this strain on EDTA mineral salt medium always stopped before the substrate was completely degraded and the remaining concentration of EDTA corresponded to the Fe(III) concentration in the medium. Nevertheless, growth of BNC1 was dependent on the presence of metal-complexed EDTA because free EDTA appeared to interact negatively with the cell walls and completely inhibited bacterial growth. Experiments with resting cells of strain BNC1 revealed that EDTA complexes with low stability constants (log K 6 14) were readily degraded. CaEDTA, BaEDTA and MgEDTA were consumed at the highest rates, followed by MnEDTA. Free EDTA was also degraded, although growth with this species was observed to slow down. Generally, chelates with stability constants exceeding 1014 were not degraded. Only ZnEDTA (log K = 16.5) was slowly oxidised [232]. Similar results were reported recently for strain DSM 9103 [250]. Complexes with stability constants (log K 6 16, i.e. the Mg2‡ , Ca2‡ , Mn2‡ complexes and free EDTA) were degraded by EDTA-grown resting cells to completion at a constant rate. For more stable EDTA chelates (with Co2‡ , Cu2‡ , Zn2‡ and Pb2‡ ), the data suggested that these complexes were not used directly but had to dissociate prior to degradation. Only CdEDTA and Fe(III)EDTA were not degraded within 48 h. In the case of CdEDTA, the toxicity of freed Cd2‡ ions most likely prevented a signi¢cant degradation of EDTA, whereas in the case of Fe(III)EDTA, a combination of metal toxicity and the very slow dissociation of the complex was probably responsible for the absence of degradation [250]. For transport in strain DSM 9103, uptake experiments with [14 C]EDTA in presence of various metal ions revealed a nearly identical pattern to that described for resting cells of strain BNC1 [231]. Free EDTA as well as chelates of Ca2‡ , Ba2‡ and Mg2‡ were readily incorporated at similar rates. Uptake of MnEDTA proceeded at an intermediate rate and all other, more stable complexes (those with Zn2‡ , Co2‡ , Cu2‡ , Ni2‡ or Fe3‡ ) were not transported in these short-term assays. In addition, Ca2‡ uptake of strain DSM 9103 was considerably increased in the presence of EDTA, indicating that Ca2‡ is probably transported together with EDTA into the cells. Cells of strain DSM 9103 seem to utilise both proposed mechanisms, i.e. intracellular precipitation and excretion of the metal, to cope with the increased incoming £ux of metal ions due to EDTA degradation. All these studies indicate that for the two strains BNC1 and DSM 9103, the EDTA chelates of Zn2‡ , Cu2‡ , Ni2‡ , Co2‡ and Fe3‡ are not directly accessible. They can only be degraded after dissociation, a process that depends on the dissociation rate constant of the complex. However, this rate constant is rather low for some of the heavy

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

metal^EDTA complexes, such as Fe(III)EDTA, NiEDTA or CuEDTA [91]. Consequently, the environmental recalcitrance of EDTA towards biological degradation might ^ at least partly ^ be attributed to the presence of species which do not favour microbial attack, in particular when considering that in a typical river system about 70% of the total EDTA is predicted to be present as stable heavy metal complexes (Fig. 3) [85]. 7.2. At the enzyme level At the enzymatic level, NTA MO from C. heintzii ATCC 29600 [216,251], NTA DH from strain TE11 [229,252], EDTA MO from strain DSM 9103 [198], cellfree extracts and EDTA MO from strain BNC1 [232,234], and ¢nally EDDS lyase from strain DSM 9103 [235] were 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 MO from strain DSM 9103 exhibited a similar behaviour. Neither of the MOs transformed the free complexing agent or the Ca2‡ chelate, whereas complexes of NTA or EDTA with Mg2‡ , Mn2‡ , Co2‡ and Zn2‡ were degraded. NTA MO attacked also Fe(III)NTA and NiNTA. The highest speci¢c transformation activity was found for the Mg2‡ complexes. Since Mg2‡ is the most abundant intracellular divalent cation [253], NTA and EDTA can be expected to be mainly present as Mg2‡ chelate in the cytoplasm. Therefore, it is not surprising that Mg complexes of both APCAs were the best substrates. All in all, no clear relationship is discernible between the stability of a NTA or EDTA complex or its dissociation rate constant and its MO-catalysed oxidation. Therefore, it was suggested that the structure of a complex might be a determining factor [198]. It should be added that for the EDTA-degrading strain BNC1, con£icting results were published. In cell-free extracts, Klu«ner and co-workers [232] found a similar pattern to that described above for the two MOs, with free EDTA and CaEDTA not being oxidised. In contrast, these two species, as well as various other EDTA complexes with di- and trivalent cations, were degraded in enzyme assays with puri¢ed EDTA MO performed by Payne and co-workers [234]. Currently, this discrepancy cannot be explained. In contrast to NTA MO and EDTA MO from DSM 9103, NTA DH accepted free NTA. Addition of divalent metal ions almost completely inhibited enzymatic NTA breakdown [252]. This inhibition was reversed upon addition of EDTA, strongly suggesting that only uncomplexed NTA is a substrate of the DH. Similarly, EDDS lyase accepted free [S,S]-EDDS as a substrate but also EDDS complexes with Mg2‡ , Ca2‡ , Ba2‡ and Mn2‡ [235]. All these complexes have stability constants lower than 109 . MnEDDS with the highest stability constant of all degradable species was trans-

97

formed at a very low rate. All other complexes which were not attacked by the lyase exhibited stability constants higher than 1010 . Hence, in the case of EDDS, the stability of a complex apparently rules the enzymatic degradability. One can speculate that during substrate-binding, the metal ion has to be exchanged or removed from the complex with the EDDS molecule. This assumption is supported by the observation that metal ions had a strong in£uence on the equilibrium between EDDS and the cleavage products which is reached in the lyase-catalysed elimination reaction (Fig. 6). In presence of metal ions (especially those forming non-degradable complexes), this equilibrium was shifted towards the educt EDDS probably because free EDDS was removed from the equilibrium by the formation of metal complexes, indicating that only free EDDS is the actual substrate of the lyase. 8. Outlook To date, the industrially most important APCAs are the synthetically produced compounds NTA, EDTA, DTPA and HEDTA. Recently, a number of new APCAs, such as PDTA, L-ADA, SDA or MGDA, have been synthesised which may have the potential to substitute the classical representatives in various applications and they are currently being tested. In addition, several naturally occurring APCAs produced by micro-organisms or plants, also with considerable potential for application, have been described in the literature. The producing organisms exploit the metal-complexing capacity of APCAs for enhancing their metal uptake, in most cases that of iron. On top of their good complexing ability, a considerable advantage of these natural APCAs is that they are probably better biodegradable than most synthetic APCAs. This is because, ¢rstly, the metal they are transporting into the cell has to be freed intracellularly and, secondly, because organisms have been exposed to them for a longer time than to the xenobiotic ones. However, to date, there is little information available on these aspects. First studies with respect to a biotechnological production of some of these APCAs have already been reported and they seem promising. Nevertheless, considerably more research is still needed before such biotechnologically produced APCAs may compete with the established and cheap synthetic APCAs NTA and EDTA. Once released into the environment, the fate of APCAs is determined either by abiotic or biotic processes (Fig. 7). For each of the di¡erent compounds, it has to be assessed individually because no general pattern applicable to all of them has emerged. The environmental fate of NTA is determined by biodegradation with abiotic processes playing a minor role for its elimination from both wastewater and ecosystems. Hence, NTA most likely does not present a major risk for the environment. In contrast, EDTA, DTPA and HEDTA seem to be ^ if at all ^ only poorly biodegradable and therefore abiotic elimination mecha-

FEMSRE 706 29-12-00

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 deacylation. 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 been demonstrated that the most important elimination process in rivers is its photolysis, which is, however, restricted to sunny days and to only one EDTA species, i.e. Fe(III)EDTA. To what extent abiotic oxidative processes are important for EDTA (and perhaps also other APCAs) breakdown in soils, especially in comparison to biodegradation, remains to be investigated. Furthermore, one should not forget that abiotic processes usually do not lead to a mineralisation of the complexing agents, but merely to a transformation into compounds which still have metal-complexing properties. Therefore, future research should also focus on the metabolites of these abiotic degradation processes such as ED3A. Especially in monitoring programs for EDTA and DTPA potential breakdown products should be included, too, for a better assessment of the pollution of a system with these xenobiotic APCAs. As mentioned above, new APCAs of both synthetic and biological origin are being tested for their

potential to replace the classical ones. Some of them appear to be easily biodegradable, at least under aerobic conditions. Though, for all of them, the information available is still too limited to conclude that they are harmless from an environmental point of view. So far, the isolation of aerobic strains growing with NTA, EDTA or [S,S]-EDDS has been successful. Virtually all strains have been found to be members of the Agrobacterium^Rhizobium group in the K-branch of Proteobacteria. A possible future isolation and taxonomic characterisation of bacterial strains growing with DTPA, HEDTA, L-ADA, SDA, ASDA or MGDA might unravel an interesting picture with respect to the evolution of microbial degradation of xenobiotic APCAs. In the absence of molecular oxygen, NTA is the only APCA for which degradation has been reported under denitrifying conditions. Interestingly, this strain is not a member of the K-branch of Proteobacteria. Some of the NTA- and EDTA-degrading strains pres-

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

ently isolated seem to exhibit an extended spectrum of APCAs they can utilise as growth substrates. One can therefore expect that testing the newly marketed chelating agents such as PDTA, L-ADA, EDDS or MGDA for degradation by NTA- or EDTA-utilising bacteria might yield interesting results. Furthermore, the fact that both the EDTA-degrading strain investigated in our laboratory and the NTA-degrading Chelatococcus strains are able to degrade not only one but two or more APCAs indicates that such pure cultures could perhaps be used for special treatment of wastewaters which contain high amounts of rather poorly biodegradable APCAs such as DTPA, HEDTA or PDTA. In case of EDTA, the application of pure cultures has already shown rather promising results in laboratory experiments [212,213], whereas large scale treatment still remains to be investigated. Due to the isolation of pure bacterial cultures, the metabolic 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 cleavage of the bond between a nitrogen atom and a carboxylic residue (an acetyl or succinyl group). So far, decarboxylation or cleavage within the ethylenediamine part of the EDTA or EDDS molecule has not been observed as ¢rst metabolic step. In this context, it is worth mentioning that the biochemistry of EDTA degradation in the Agrobacterium strain which grows only with iron(III)-complexed EDTA [190], a complex that is not touched by the strains investigated so far, has not been studied yet. Perhaps, the co-ordinated iron ion plays a role in the enzymatic breakdown of EDTA in this bacterium. Of course, for the characterisation of degradative pathways for other APCAs such as DTPA or the amino acid derivatives L-ADA, SDA, ASDA or MGDA, pure cultures able to degrade these chemicals should be available. However, one can speculate that the aerobic breakdown will also include the attack of an MO on the compounds as seen for NTA and EDTA degradation in strains C. heintzii, BNC1 and DSM 9103. This would lead to the removal of an acetyl group. In fact, EDTA MO has been shown not only to catalyse the deacetylation of EDTA but also that of DTPA and HEDTA. Therefore, one can speculate that L-ADA, SDA, ASDA or MGDA should all be accepted as substrates by both NTA MO and NTA DH due to the structural resemblance of these compounds with NTA. In analogy to the NTA metabolism for the amino acid derivatives, a second acetyl group can be supposed to be cleaved o¡ in a reaction catalysed by a DH and generating the corresponding amino acid. From a number of studies, it has become clear that the speciation of APCA molecules greatly in£uences their environmental fate. But especially in case of biological elimination, not enough information is currently available to suggest how exactly this process is a¡ected by speciation. So far, it has not been shown for any APCA-degrading bacterium which species of the complexing compound is/

99

are really taken up by the cells. This information, however, seems essential for understanding why the biodegradability of di¡erent APCAs varies so tremendously. If only free APCA molecules or rather unstable species are transported into the cells, then this might explain why APCAs such as EDTA or DTPA forming very stable complexes (especially with heavy metals) are not as well biodegradable as NTA. This could also explain the positive e¡ect of a pH rise on the microbial elimination of EDTA as it was observed in a sewage treatment plant because it is known that heavy metal^EDTA complexes are destabilised at higher pH values. All in all, the enormous in£uence speciation has on the fate of APCAs in nature shows clearly the necessity to include all important species of new APCAs in degradation and toxicity tests in order to be really able to predict and judge the compound's environmental behaviour. Acknowledgements The authors would like to thank Kay Fox, Laura Sigg and Bernd Nowack for careful reading of and valuable comments on this manuscript. TE would like to express his thanks to Hansueli Weilenmann for the fruitful collaboration over the years (not only in the ¢eld of complexing agent biodegradation) as well as to all the research students that have added mosaic stone by mosaic stone without which it would have been impossible to get to the present level of understanding. Furthermore, the generous ¢nancial support from the Swiss National Science Foundation, The Research Commission of the Swiss Federal Institute of Technology Zu«rich, Lever Switzerland and Unilever Reserach Laboratories Port Sunlight (UK), the Swiss Agency for the Protection of the Environment (BUWAL), and last, but not least, of the Swiss Federal Institute for Environmental Science and Technology is gratefully acknowledged.

References [1] Bell, C.F. (1977) Principles and Applications of Metal Chelation. Clarendon Press, Oxford. [2] Wilkinson, G. (1987) Comprehensive Coordination Chemistry. Pergamon Press, Oxford. [3] Egli, T. (1988) (An)aerobic breakdown of chelating agents used in household detergents. Microbiol. Sci. 5, 36^41. [4] Anderson, R.L., Bishop, E.B. and Campbell, R.L. (1985) A review of the environmental and mammalian toxicology of nitrilotriacetic acid. Crit. Rev. Toxicol. 15, 1^102. [5] Egli, T., Bally, M. and Uetz, T. (1990) Microbial degradation of chelating agents used in detergents with special reference to nitrilotriacetic acid (NTA). Biodegradation 1, 121^132. [6] Egli, T. (1994) Biochemistry and physiology of the degradation of nitrilotriacetic acid and other metal complexing agents. In: Biochemistry of Microbial Degradation (Ratledge, C., Ed.), pp. 179^195. Kluwer Academic Publishers, Dordrecht.

FEMSRE 706 29-12-00

100

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

[7] Kari, F.G. and Giger, W. (1995) Modeling the photochemical degradation of ethylenediaminetetraacetate in the river Glatt. Environ. Sci. Technol. 29, 2814^2827. ë ber dem Ammoniaktypus angeho«rige Sa«uren. [8] Heintz, W. (1862) U Ann. Chem. Pharm. 122, 257^294. [9] Pottho¡-Karl, B. (1994) Neue biologisch abbaubare Komplexbildner. ë le Fette Wachse 120, 104^109. Seifen O [10] Parker, B.A. and Crudden, J.J. (1996) The commercial synthesis and characterization of novel multifunctional sufactant chelates. Abstract presented at the 4th World Surfactant Conference, pp. 446^460. Barcelona. [11] Nishikiori, T., Okuyama, A., Naganawa, T., Takita, T., Hamada, M., Takeuchi, T., Aoyagi, T. and Umezawa, H. (1984) Production by actinomycetes of (S,S)-N,NP-ethylenediamine-disuccinic acid, an inhibitor of phospholipase C. J. Antibiot. 37, 426^427. [12] Cebulla, I. (1995) Gewinnung komplexbildender Substanzen mittels Amycolatopsis orientalis. Ph.D. Thesis, Eberhard-Karls-Universita«t Tu«bingen, Tu«bingen. [13] Cebulla, I., Harder, M., Theobald, U. and Za«hner, H. (1996) Studies on the production of ethylene-diamine-disuccinic acid by Amycolatopsis orientalis. Poster presented at the 55th Annual Meeting of the Swiss Society of Microbiology, Bern. [14] Majer, J., Springer, V. and Kopecka, B. (1966) New complexones. VIII. Ethylenediamino-N,NP-disuccinic acid and investigation of its heavy metal complexes by spectrophotometry. Chem. Zvesti 20, 414^ 422. [15] Gorelov, I.P., Samsonov, A.P., Nikol'skii, V.M., Babich, V.A., Svetogorov, Y.E., Smirnova, T.I., Malakhaev, E.D., Kozlov, Y.M. and Kapustnikov, A.I. (1979) Synthesis and complex-forming properties of complexons derived from dicarboxylic acids. V. Synthesis of complexons derived from succinic acid. Zhurnal Obshchei Khimii 49, 659^663. [16] Hartmann, F.A. and Perkins, C.M. (1987) Detergent composition containing ethylenediamine-N,NP-disuccinic acid. US patent 4,704,233. [17] Neal, J.A. and Rose, N.J. (1968) Stereospeci¢c ligands and their complexes. I. A cobalt(III)complex of ethylenediaminedisuccinic acid. Inorg. Chem. 7, 2405^2412. [18] Zwicker, N., Theobald, U., Za«hner, H. and Fiedler, H.-P. (1997) Optimization of fermentation conditions for the production of ethylene-diamine-disuccinic acid by Amycolatopsis orientalis. J. Ind. Microbiol. Biotechnol. 19, 280^285. [19] Smith, M.J. and Neilands, J.B. (1984) Rhizobactin, a siderophore from Rhizobium meliloti. J. Plant Nutr. 7, 449^458. [20] Smith, A.W., Shoolery, J.N., Schwyn, B., Holden, I. and Neilands, J.B. (1985) Rhizobactin, a structurally novel siderophore from Rhizobium meliloti. J. Am. Chem. Soc. 107, 1739^1743. [21] Drechsel, H., Metzger, J., Freund, S., Jung, G., Boelaert, J.R. and Winkelmann, G. (1991) Rhizoferrin ^ a novel siderophore from the fungus Rhizopus microsporus var. rhizopodiformis. BioMetals 4, 238^ 243. [22] Thieken, A. and Winkelmann, G. (1992) Rhizoferrin: A complexone type siderophore of the Mucorales and Entomophtorales (Zygomycetes). FEMS Microbiol. Lett. 94, 37^42. [23] Winkelmann, G. (1993) Kinetics, energetics and mechanisms of siderophore iron transports in fungi. In: Iron Chelation in Plants and Soil Microorganisms (Barton, L.L., Ed.), pp. 220^239. Academic Press, Inc., London. [24] Carrano, C.J., Drechsel, H., Kaiser, D., Jung, G., Matzanke, B., Winkelmann, G., Rochel, N. and Albrecht-Gary, A.M. (1996) Coordination chemistry of the carboxylate type siderophore rhizoferrin: The iron(III) complex and its metal analogs. Inorg. Chem. 35, 6429^ 6436. [25] Carrano, C.J., Thieken, A. and Winkelmann, G. (1996) Speci¢city and mechanism of rhizoferrin-mediated metal iron uptake. BioMetals 9, 185^189. [26] Drechsel, H., Freund, S., Nicholson, G., Haag, H., Jung, O., Za«hner,

[27]

[28]

[29]

[30]

[31]

[32]

[33]

[34]

[35]

[36] [37]

[38]

[39]

[40]

[41]

[42]

[43]

[44]

[45]

H. and Jung, G. (1993) Puri¢cation and chemical characterization of staphyloferrin B, a hydrophilic siderophore from staphylococci. BioMetals 6, 185^192. Ku«hn, S., Braun, V. and Ko«ster, W. (1996) Ferric rhizoferrin uptake into Morganella morganii: Characterization of genes involved in the uptake of a polyhydroxycarboxylate siderophore. J. Bacteriol. 178, 496^504. Drechsel, H., Tschierske, M., Thieken, A., Jung, C., Za«hner, H. and Winkelmann, G. (1995) The carboxylate type siderophore rhizoferrin and its analogs produced by direct fermentation. J. Ind. Microbiol. 14, 105^112. Meiwes, J., Fiedler, H.-P., Haag, H., Za«hner, H., Konetschny-Rapp, S. and Jung, G. (1990) Isolation and characterization of staphyloferrin A, a compound with siderophore activity from Staphylococcus hyicus DSM 20459. FEMS Microbiol. Lett. 67, 201^206. Konetschny-Rapp, S., Jung, G., Meiwes, J. and Za«hner, H. (1990) Staphyloferrin A: a structurally new siderophore from staphylococci. FEBS Eur. J. Biochem. 191, 65^74. Haag, H., Fiedler, H.-P., Meiwes, J., Drechsler, H., Jung, G. and Za«hner, H. (1994) Isolation and biological characterization of staphyloferrin B, a compound with siderophore activity from staphylococci. FEMS Microbiol. Lett. 115, 125^130. Budesinsky, M., Budzikiewics, H., Prochazka, Z., Ripperger, H., Ro«mer, A., Scholz, G. and Schreiber, K. (1980) Nicotianamine, a possible phytosiderophore of general occurrence. Phytochemistry 19, 2295^2297. Takagi, S.-I. (1993) Production of phytosiderophores. In: Iron Chelation in Plants and Soil Microorganisms (Barton, L.L. and Hemming, B.C., Eds.), pp. 111^131. Academic Press, Inc., London. Cakmak, I., Ozturk, L., Karanlik, S., Marschner, H. and Ekiz, H. (1996) Zink-e¤cient wild grasses enhance release of phytosiderophores under zinc de¢ciency. J. Plant Nutr. 19, 551^563. Kawai, S., Takagi, S. and Ojima, K. (1992) Application of phytosiderophore to plant cell cultures and production of phytosiderophore by iron de¢ciency stressed plant cell cultures. J. Plant Nutr. 15, 1613^ 1624. Ro«mpp, H. (1950) Chemie Lexikon. Franckh'sche Verlagshandlung, Stuttgart. Wolf, K. and Gilbert, P.A. (1992) EDTA-ethylenediaminetetraacetic acid. In: The Handbook of Environmental Chemistry, Vol. 3 (Hutzinger, O., Ed.), pp. 241^259. Springer, Berlin. Klopp, R. and Pa«tsch, B. (1994) Organische Komplexbildner in Abwasser, Ober£a«chenwasser und Trinkwasser, dargestellt am Beispiel der Ruhr. Wasser Boden 8, 32^37. Sacher, F., Lochow, E. and Brauch, H.-J. (1998) Synthetic organic complexing agents ^ analysis and occurrence in surface waters. Vom Wasser 90, 31^41. Svenson, A., Kaj, L. and Bjo«rndal, H. (1989) Aqueous photolysis of the iron(III) complexes of NTA, EDTA and DTPA. Chemosphere 18, 1805^1808. Zhao, F., Yang, J. and Scho«neich, C. (1996) E¡ects of polyaminocarboxylate metal chelators on iron-thiolate induced oxidation of methionine- and histidine-containing peptides. Pharmacol. Res. 13, 931^938. Means, J.L. and Alexander, C.A. (1981) The environmental biochemistry of chelating agents and recommondations for the disposal of chelated radioactive wastes. Nucl. Chem. Waste Manag. 2, 183^196. Toste, A.P. and Lechner-Fish, T.J. (1993) Chemo-degradation of chelating and complexing agents in a simulated, mixed nuclear waste. Waste Manag. 13, 237^244. Macaskie, L. (1991) The application of biotechnology to the treatment of wastes produced from the nuclear fuel cycle: Biodegradation and bioaccumulation as means of radionuclide-containig streams. Crit. Rev. Biotechnol. 11, 41^112. Means, J.L., Crerar, D.A. and Duguid, J.O. (1978) Migration of radioactive wastes: Radionuclide mobilization by complexing agents. Science 200, 1477^1480.

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106 [46] Elliott, H.A. and Brown, G.A. (1989) Comparative evaluation of NTA and EDTA for extractive decontamination of Pb-polluted soils. Water Air Soil Pollut. 45, 361^369. [47] Brown, G.A. and Elliott, H.A. (1992) In£uence of electrolytes on EDTA extraction of Pb from polluted soil. Water Air Soil Pollut. 62, 157^165. [48] Hong, J. and Pintauro, P.N. (1994) Desorption^complexation^dissolution characteristics of adsorbed cadmium from kaolin by chelators. Water Air Soil Pollut. 86, 35^50. [49] Yu, K.-C., Ho, S.-T., Tsai, L.-J., Chang, J.-S. and Lee, S.-Z. (1996) Remobilization of zinc from Ell-Ren river sediment fractions a¡ected by EDTA, DTPA and EGTA. Water Sci. Technol. 34, 125^ 132. [50] Huang, J.W.W., Chen, J.J., Berti, W.R. and Cunningham, S.D. (1997) Phytoremediation of lead-contaminated soils : Role of synthetic chelates in lead phytoextraction. Environ. Sci. Technol. 31, 800^805. [51] Rubin, M. and Martell, A.E. (1980) The implication of trace metal^ nitrilotriacetic acid speciation on its environmental impact and toxicology. Biol. Trace Elem. Res. 2, 1^19. [52] Wallace, A., Wallace, G.A. and Cha, J.W. (1992) Some modi¢cations in trace metal toxicities and de¢ciencies in plants resulting from interactions with other elements and chelating agents ^ the special case of iron. J. Plant Nutr. 15, 1589^1598. [53] Deacon, M.S.M.R. and Tuinstra, L.G.M.T. (1994) Chromatographic separation of metal chelates present in commercial fertilizers. II. Development of an ion-pair chromatographic separation and simultaneous determination of the Fe(III) chelates of EDTA, DTPA, HEEDTA, EDDHA and EDDHMA and the Cu(II), Zn(II) and Mn(II) chelates of EDTA. J. Chromatogr. 659, 349^357. [54] Wei, N., Crescoulo, P.P. and LeClair, B.P. (1979) Impact of nitrilotriacetic acid (NTA) on an activated sludge plant ^ a ¢eld study Project No. 71-3-3. Environmental Protection Service Environment Canada. [55] Alder, A.C., Siegrist, H., Gujer, W. and Giger, W. (1990) Behaviour of NTA and EDTA in biological wastewater treatment. Water Res. 24, 733^742. [56] Kari, F.G. and Giger, W. (1996) Speciation and fate of ethylenediaminetetraacetate (EDTA) in municipal wastewater treatment. Water Res. 30, 122^134. [57] Alder, A.C., Siegrist, H., Fent, K., Egli, T., Molnar, E., Poiger, T., Scha¡ner, C. and Giger, W. (1997) The fate of organic pollutants in wastewater and sludge treatment : Signi¢cant processes and impact of compound properties. Chimia 51, 922^928. [58] de Oude, I.N.T. (1984) NTA-Monitoring ^ Organisation und Erfahrungen von Kanada, USA und den Niederlanden. In: NTA: Studie u«ber die aquatische Umweltvertra«glichkeit von Nitrilotriacetat (NTA) (Bernhardt, H., Ed.), pp. 413^422. Verlag Hans Richarz, Sankt Augustin. [59] Woodiwiss, C.R., Walker, R.D. and Brownridge, F.A. (1979) Concentration of nitrilotriacetate and certain metals in Canadian wastewaters and streams: 1971^1975. Water Res. 13, 599^612. [60] Ernst, W. and Kleiser, H.H. (1984) Vorkommen, Verhalten und Auswirkungen von NTA im marinen Bereich. In: NTA: Studie u«ber die aquatische Umweltvertra«glichkeit von Nitrilotriacetat (NTA) (Bernhardt, H., Ed.), pp. 237^250. Verlag Hans Richarz, Sankt Augustin. [61] Houriet, J.-P. (1996) NTA dans les eaux. Cahier de l'environnement 264. O¤ce fe¨de¨ral de l'environnement, des foreªts et du paysage (OFEFP), Bern. [62] Kari, F.G. (1994) Umweltverhalten von Ethylendiamintetraacetat (EDTA) unter spezieller Beru«cksichtigung des photochemischen Abbaus. Ph.D. Thesis No 10698, Swiss Federal Institute of Technology, Zu«rich. [63] Ko«nen, I. (1997) Bestimmung von EDTA-Ersatzsto¡en auf Aminopolycarbonsa«urebasis, Vol. 159. Gesellschaft zur Fo«rderung der Siedlungswasserwirtschaft and der RWTH Aachen e.V., Aachen.

101

[64] Nowack, B., Kari, F.G., Hilger, S.U. and Sigg, L. (1996) Determination of dissolved and adsorbed EDTA species in water and sediments by HPLC. Anal. Chem. 68, 561^566. [65] Sillanpa«a, M., Vickackaite, V., Niinisto«, L. and Sihvonen, M.L. (1997) Distribution and transportation of ethylenediaminetetraacetic acid and diethylenetriaminepentaacetic acid in lake water and sediment. Chemosphere 35, 2797^2805. [66] Sillanpa«a, M. and Oikari, A. (1996) Transportation of complexing agents released by pulp and paper industry: a Finnish lake case. Toxicol. Environ. Chem. 57, 79^91. [67] Dietz, F. (1987) Neue Messergebnisse u«ber die Belastung von Trinkwasser mit EDTA. gwf Wasser Abwasser 128, 286^288. [68] Grischek, T., Neitzel, P., Andrusch, T., Lagois, U. and Nestler, W. (1997) Fate of EDTA during in¢ltration of Elbe river water and identi¢cation of in¢ltrating river water in the aquifer. Vom Wasser 89, 261^282. [69] Lindner, K., Knepper, T.P., Karrenbrock, F., Ro«rden, O., Brauch, H.-J., Lange, F.T. and Sacher, F. (1996) Erfassung und Identi¢zierung von trinkwasserga«ngigen Einzelsubstanzen in Abwa«ssern und im Rhein, Vol. 1. IAWR, Ko«ln. [70] Toste, A.P., Osborn, B.C., Polach, K.J. and Lechner-Fish, T.J. (1995) Organic analyses of an actual and simulated mixed waste: Hanford's organic complexant waste revisited. J. Radioanal. Nucl. Chem. 194, 25^34. [71] Buchberger, W., Haddad, P.R. and Alexander, P.W. (1991) Separation of metal complexes of ethylenediaminetetraacetic acid in environmental water samples by ion chromatography with UV and potentiometric detection. J. Chromatogr. 558, 181^186. [72] Buchberger, W. and Mu«lleder, S. (1995) Determination of chelating agents and metal chelates by capillary zone electrophoresis. Microchim. Acta 119, 103^111. [73] Bu«rgisser, C.S. and Stone, A.T. (1997) Determination of EDTA, NTA, and other amino carboxylic acids and their Co(II) and Co(III) complexes by capillary electrophoresis. Environ. Sci. Technol. 31, 2656^2664. [74] Campos, M.L.A.M. and Van den Berg, C.M.G. (1994) Determination of copper complexation in sea water by cathodic stripping voltammetry and ligand competition with salicylaldoxim. Anal. Chim. Acta 284, 481^496. [75] Donat, J.R., Lao, K.A. and Bruland, K.W. (1994) Speciation of dissolved copper and nickel in South San Francisco Bay: a multimethod approach. Anal. Chim. Acta 184, 547^571. [76] Luther, G.W., Nuzzio, D.B. and Wu, J. (1994) Speciation of manganese in Chesapeake Bay waters by voltametric methods. Anal. Chim. Acta 284, 473^480. [77] Mackey, D.J. and Zirino, A. (1994) Comments on trace metal speciation in seawater or do onions grow in the sea? Anal. Chim. Acta 284, 635^647. [78] Xue, H.B. and Sigg, L. (1993) Free cupric ion concentration and Cu(II) speciation in a eutrophic lake. Limnol. Oceanogr. 38, 1200^ 1213. [79] Xue, H.B. and Sigg, L. (1994) Zinc speciation in lake waters and its determination by ligand exchange with EDTA and di¡erential pulse anodic stripping voltammetry. Anal. Chim. Acta 284, 505^515. [80] Qian, J., Xue, H.B., Sigg, L. and Albrecht, A. (1998) Complexation of cobalt by natural ligands in freshwater. Environ. Sci. Technol. 32, 2043^2050. [81] Xue, H.B. and Sigg, L. (1998) Cadmium speciation and complexation by natural organic ligands in fresh water. Anal. Chim. Acta 363, 249^ 259. [82] Hering, J.G. and Morel, F.M.M. (1988) Kinetics of trace metal complexation: role of alkaline-earth metals. Environ. Sci. Technol. 22, 1469^1478. [83] Hering, J.G. and Morel, F.M.M. (1990) Kinetics of trace metal complexation: implication for metal reactivity in natural waters. In: Aquatic Chemical Kinetics (Stumm, W., Ed.), pp. 145^171. John Wiley and Sons, New York.

FEMSRE 706 29-12-00

102

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

[84] Xue, H.B., Sigg, L. and Kari, F.G. (1995) Speciation of EDTA in natural waters: Exchange kinetics of Fe-EDTA in river water. Environ. Sci. Technol. 29, 59^68. [85] Nowack, B., Xue, H. and Sigg, L. (1997) In£uence of natural and anthropogenic ligands on metal transport during in¢ltration of river water to groundwater. Environ. Sci. Technol. 31, 866^872. [86] Nowack, B., 1996. Behaviour of EDTA in groundwater ^ a study of the surface reactions of metal^EDTA complexes. Ph.D. Thesis No. 11392, Swiss Federal Institute of Technology, Zu«rich. ë kologisch relevante [87] Scho«berl, P., Huber, M. and Huber, L. (1988) O Daten von nichttensidischen Inhaltssto¡en in Wasch- und Reinigungsmitteln. Tens. Surfactant Deterg. 25, 99^107. [88] van Dam, R.A., Barry, M.J., Ahokas, J.T. and Holdway, D.A. (1996) Comparative acute and chronic toxicity of diethylenetriamine pentaacetic acid (DTPA) and ferric-complexed DTPA to Daphnia carinata. Arch. Environ. Contam. Toxicol. 31, 433^443. [89] Allen, H.E. (1983) Potential for metal mobilization by synthetic organic chelating agents ^ a case study. Presented at the International Conference `Heavy metals in the environment', Heidelberg. [90] Twachtmann, U., Petrick, S., Merz, W. and Metzger, J.W. (1998) Zum Ein£uM umweltrelevanter Konzentrationen des Komplexbildners EDTA auf die Remobilisierung von Schwermetallen im Belebungsverfahren. Vom Wasser 91, 101^120. [91] Nowack, B. and Sigg, L. (1997) Dissolution of Fe(III)(hydr)oxides by metal^EDTA complexes. Geochim. Cosmochim. Acta 61, 951^ 963. [92] Davis, J.A., Kent, D.B., Rea, B.A., Maest, A.S. and Garabedian, S.P. (1993) In£uence of redox environment and aqueous speciation on metal transport in groundwater : preliminary results of trace injection studies. In: Metals in Groundwater (Allen, H.E., Perdue, E.M. and Brown, D.S., Eds.). Lewis Publishers, Chelsea. [93] Langford, C.H., Wingham, M. and Sastri, V.S. (1973) Ligand photooxidation in copper(II) complexes of nitrilotriacetic acid. Environ. Sci. Technol. 7, 820^822. [94] Stolzberg, R.J. and Hume, D.N. (1975) Rapid formation of iminodiacetate from photochemical degradation of Fe(III) nitrilotriacetate solutions. Environ. Sci. Technol. 9, 654^656. [95] Mailhot, G., Bordes, A.-L. and Bolte, M. (1995) Iminodiacetic acid degradation photoinduced by complexation with monometallic (iron(III)) and bimetallic systems (iron(III) and copper(II)). Chemosphere 30, 1729^1737. [96] Lockhart Jr., H.B. and Blakeley, R.V. (1975) Aerobic photodegradation of Fe(III)-(ethylenedinitrilo)tetraacetate (ferric EDTA). Environ. Sci. Technol. 9, 1035^1038. [97] Natarajan, P. and Endicott, J.F. (1973) Photoredox behavior of transition metal^ethylenediaminetetraacetate complexes. A comparison of some group VIII metals. J. Phys. Chem. 77, 2049^2054. [98] Karametaxas, G., Hug, S.J. and Sulzberger, B. (1995) Photodegradation of EDTA in the presence of lepidocrocite. Environ. Sci. Technol. 29, 2992^3000. [99] Frank, R. and Rau, H. (1989) Photochemical transformation in aqueous solution and possible environmental fate of ethylenediaminetetraacetatic acid (EDTA). Ecotox. Environ. Saf. 19, 55^63. [100] Kari, F.G., Hilger, S. and Canonica, S. (1995) Determination of the reaction quantum yield for the photochemical degradation of Fe(III)-EDTA : Implications for the environmental fate of EDTA in surface waters. Environ. Sci. Technol. 29, 1008^1017. [101] Nowack, B. and Baumann, U. (1998) Biodegradation of the photolysis products of Fe(III)EDTA. Acta Hydrochim. Hydrobiol. 26, 104^108. [102] Gardiner, J. (1976) Complexation of trace metals by ethylenediaminetetraacetic acid (EDTA) in natural waters. Water Res. 10, 507^ 514. [103] Nowack, B. and Sigg, L. (1996) Adsorption of EDTA and metal^ EDTA complexes onto goethite. J. Colloid Interface Sci. 177, 106^ 121. [104] Bowers, A.R. and Huang, C.P. (1986) Adsorption characteristics of

[105]

[106]

[107]

[108]

[109]

[110] [111] [112] [113]

[114]

[115]

[116]

[117] [118]

[119] [120]

[121]

[122]

[123]

[124]

[125]

metal^EDTA complexes onto hydrous oxides. J. Colloid Interface Sci. 110, 575^590. Ulrich, M. (1991) Modeling of chemicals in lakes ^ development and application of user-friendly simulation software (MASAS and CHEMSEE). Ph.D. Thesis No 9632, Swiss Federal Institute of Technology, Zu«rich. Klewicki, J.K. and Morgan, J.J. (1998) Kinetic behavior of Mn(III) complexes of pyrophosphate, EDTA, and citrate. Environ. Sci. Technol. 32, 2916^2922. McArdell, C.S., Stone, A.T. and Tian, J. (1998) Reaction of EDTA and related aminopolycarboxylate chelating agents with Co(III)OOH (Heterogenite) and Mn(III)OOH (Manganite). Environ. Sci. Technol. 32, 2923^2930. Norvell, W.A. (1991) Reactions of metal chelates in soils and nutrient solutions. In: Micronutrients in agriculture (Mortvedt, J.J., Cox, F.R., Shuman, L.M. and Welch, R.M., Eds.). Soil Sci. Soc. Am., Inc., Madison, WI. Shumate, K.S., Thompson, J.E., Brookhart, J.B. and Dean, C.L. (1970) NTA removal by activated sludge ^ ¢eld study. J. Water Pollut. Control Fed. 42, 631^640. Bouveng, H.O., Salyom, P. and Werner, J. (1970) Degradation of NTA in a trickling ¢lter and an oxidation pond. Vatten 4, 389^402. Gundernatsch, H. (1974) Biologischer Abbau von Nitrilotriessigsa«ure. gwf Wasser Abwasser 115, 418^421. Renn, E. (1974) Biodegradation of NTA detergents in wastewater treatment systems. J. Water Pollut. Control Fed. 46, 2363^2371. Cleasby, J.L., Hubly, D.W., Ladd, T.A. and Schon, E.A. (1974) Trickling ¢ltration of a waste containing NTA. J. Water Pollut. Control Fed. 46, 1873^1887. Shannon, E. (1975) E¡ects of detergent formulation on wastewater characteristics and treatment. J. Water Pollut. Control Fed. 47, 2371^2383. Giger, W., Brunner, P.H., Ahel, M., McEvoy, J., Marcomini, A. and Scha¡ner, C. (1987) Organische Waschmittelinhaltsto¡e und deren Abbauprodukte in Abwasser und Kla«rschlamm. Gas-Wasser-Abwasser 67, 111^122. Siegrist, H., Alder, A., Gujer, W. and Giger, W. (1989) Behaviour and modelling of NTA degradation in activated sludge systems. Water Sci. Technol. 21, 315^324. Klein, S.A. (1974) NTA removal in septic tank and oxidation pond systems. J. Water Pollut. Control Fed. 46, 78^88. Kirk, P.W.W., Lester, J.N. and Perry, R. (1982) The behavior of nitrilotriacetic acid during the anaerobic digestion of sewage sludge. Water Res. 16, 973. Moore, L. and Barth, E.F. (1976) Degradation of NTA during anaerobic digestion. J. Water Pollut. Control Fed. 48, 2406^2414. Bernhardt, H., Berth, W., Fo«rster, U., Hamm, A., Janicke, W., Kandler, J., Kanowski, S., Kleiser, H.H., Koppe, P., Opgenorth, H.J., Reichert, J.K. and Stehfest, H. (1984) NTA: Studie u«ber die aquatische Umweltvertra«glichkeit von Nitrilotriacetat (NTA). Verlag Hans Richarz, Sankt Augustin. Kirk, P.W.W., Lester, J.N. and Perry, R. (1983) Amendability of nitrilotriacetic acid to biodegradation in a marine simulation. Mar. Pollut. Bull. 14, 88^93. Bartholomew, G.W. and Pfaender, F.K. (1983) In£uence of spatial and temporal variations on organic pollutant biodegradation rates in an estuarine environment. Appl. Environ. Microbiol. 45, 103^ 109. Pfaender, F.K., Shimp, R.J. and Larson, R.J. (1985) Adaptation of estuarine ecosystems to the biodegradation of nitrilotriacetic acid : E¡ects of preexposure. Environ. Toxicol. Chem. 4, 587^593. Palumbo, A.V., Pfaender, F.K. and Paerl, H.W. (1988) Biodegradation of NTA and m-cresol in coastal environments. Environ. Toxicol. Chem. 7, 573^585. Janicke, W., Fischer, W.K., Gudernatsch, H., Gu«nther, K.O., Opgenorth, H.-J., de Oude, N.T. and Wunderlich, M. (1984) Grundlagen des Abbaus und der Elimination von Nitrilotriessigsa«ure

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

[126]

[127]

[128] [129]

[130]

[131]

[132] [133]

[134]

[135] [136]

[137] [138]

[139]

[140]

[141]

[142]

[143]

[144] [145]

[146]

(NTA), NTA-Metall-Komplexen und Folgeprodukten (Mechanismen, Kinetik). In: NTA: Studie u«ber die aquatische Umweltvertra«glichkeit von Nitrilotriacetat (NTA) (Bernhardt, H., Ed.), pp. 139^165. Verlag Hans Richarz, Sankt Augustin. Larson, R.J. and Davidson, D.H. (1982) Acclimation to and biodegradation of nitrilotriacetate (NTA) at trace concentrations in natural waters. Water Res. 16, 1597^1604. Bott, T.L., Patrick, R., Larson, R. and Rhyne, C. (1978) The e¡ect of nitrilotriacetate (NTA) on the structure and functioning of aquatic communities in streams. Rep. to U.S. EPA ^ ERL, Duluth Minn. Contract No. R-801951, Environmental Protection Agency. Giger, W. (1996) Micropollutants in the environment. EAWAG News 40E, 3^7. Larson, R.J., Clickmaillie, G.G. and Van Belle, L. (1981) E¡ect of temperature and dissolved oxygen on biodegradation of nitrilotriacetate. Water Res. 15, 615^620. Kuhn, E., van Loosdrecht, M., Giger, W. and Schwarzenbach, R.P. (1987) Microbial degradation of nitrilotriacetate (NTA) during river water/groundwater in¢ltration: laboratory column studies. Water Res. 10, 1237^1248. Ahel, M., Scha¡ner, C. and Giger, W. (1996) Behaviour of alkylphenol polyethoxylate surfactants in the aquatic environment. 3. Occurrence and elimination of their persistent metabolites during in¢ltration of river water to groundwater. Water Res. 30, 37^46. Tiedje, J.M. and Mason, B.B. (1974) Biodegradation of nitrilotriacetate (NTA) in soils. Soil Sci. Am. Proc. 38, 278^283. Ward, T.E. (1985) Aerobic and anaerobic biodegradation of nitrilotriacetate in subsurface soils. Ecotoxicol. Environ. Saf. 11, 112^ 125. Shimp, R.-J., Lapsins, E.-V. and Ventullo, R.-M. (1994) Chemical fate and transport in a domestic septic system: Biodegradation of linear alkylbenzene sulfonate (LAS) and nitrilotriacetic acid (NTA). Environ. Toxicol. Chem. 13, 205^212. Tabatabai, M.A. and Bremner, J.M. (1975) Decomposition of nitrilotriacetate (NTA) in soils. Soil Biol. Biochem. 7, 103^106. Lahl, U. and Burbaum, H. (1988) Einzelsto¡analysen im Zu- und Ablauf einer kommunalen Kla«ranlage. Korresp. Abwasser 35, 360^ 364. Nowack, B. (1998) The behaviour of phosphonates in wastewater treatment plants of Switzerland. Water Res. 32, 1271^1279. Boatman, R.J., Cunninghamm, S.L. and Ziegler, D.A. (1986) A method for measuring the biodegradation of organic chemicals. Environ. Toxicol. Chem. 5, 233^243. Madsen, E.L. and Alexander, M. (1985) E¡ects of chemical speciation on the mineralization of organic compounds by microorganism. Appl. Environ. Microbiol. 50, 342^349. Hinck, M.L., Ferguson, J. and Puhaakka, J. (1997) Resistance of EDTA and DTPA to aerobic biodegradation. Water Sci. Technol. 35, 25^31. Takahashi, R., Fujimoto, N., Suzuki, M. and Endo, T. (1997) Biodegradabilities of ethylenediamine-N,NP-disuccinic acid (EDDS) and other chelating agents. Biosci. Biotechnol. Biochem. 61, 1957^ 1959. Kaluza, U., Klingelho«fer, P. and Taeger, K. (1998) Microbial degradation of EDTA in an industrial wastewater treatment plant. Water Res. 32, 2843^2845. Belly, R.T., Lau¡, J.J. and Goodhue, C.T. (1975) Degradation of ethylenediaminetetraacetic acid by microbial populations from an aerated lagoon. Appl. Microbiol. 29, 787^794. Gschwind, N. (1992) Biologischer Abbau von EDTA in einem Modellabwasser. gwf Wasser Abwasser 133, 546^549. van Ginkel, C.G., Vandenbroucke, K.L. and Stroo, C.A. (1997) Biological removal of EDTA in conventional activated-sludge plants operated under alkaline conditions. Bioresour. Technol. 59, 151^ 155. Virtaphoja, J. and Ale¨n, R. (1997) Accelerated biodegradation of EDTA in a conventional activated sludge plant under alkaline con-

[147]

[148] [149]

[150] [151]

[152]

[153] [154]

[155]

[156]

[157] [158] [159]

[160] [161]

[162]

[163]

[164]

[165]

[166]

[167] [168]

103

ditions. Contribution presented at the Proceedings TAPPI 1997 Environmental Conference, pp. 991^997. Minneapolis, MN. Langi, A., Priha, M. and Tapanila, T. (1997) Environmental fate of complexing agents in pulp and paper mill e¥uents. Contribution presented at the International conference on environmental fate and e¡ects of pulp and paper mill e¥uents, Rotorua. van Ginkel, C.G. and Boelema, E. (1999) Microbial degradation of alkylene amine acetates. US patent 5,965,024. Thomas, R.A.P., Lawlor, K., Bailey, M. and Macaskie, L.E. (1998) Biodegradation of metal^EDTA complexes by an enriched microbial population. Appl. Environ. Microbiol. 64, 1319^1322. Virtapohja, J. and Ale¨n, R. (1999) Behaviour of EDTA in marine microcosms. Chemosphere 38, 143^154. Stumpf, M., Ternes, T.A., Schuppert, B., Haberer, K., Ho¡mann, P. and Ortner, H.M. (1996) Sorption und Abbau von NTA, EDTA und DTPA wa«hrend der Bodenpassage. Vom Wasser 86, 157^ 171. Allard, A.-S., Renberg, L. and Neilson, A.H. (1996) Absence of 14 CO2 evolution from 14 C-labelled EDTA and DTPA and the sediment/water partition ratio. Chemosphere 33, 577^583. Tiedje, J.M. (1975) Microbial degradation of ethylenediaminetetraacetate in soils and sediments. Appl. Microbiol. 30, 327^329. Tiedje, J.M. (1977) In£uence of environmental parameters on EDTA biodegradation in soils and sediments. J. Environ. Qual. 6, 21^26. Means, J.L., Kucak, T. and Crerar, D.A. (1980) Relative degradation rates of NTA, EDTA and DTPA and environmental implications. Environ. Pollut. (Series B) 1, 45^60. Bolton Jr., H. (1993) Biodegradation of synthetic chelates in subsurface sediments from the Southeast coastal plain. J. Environ. Qual. 22, 125^132. Sillanpa«a, M. (1996) Complexing agents in waste water e¥uents of six Finnish pulp and paper mills. Chemosphere 33, 293^302. Nispel, F., Baumann, W. and Hardes, G. (1990) Abbauversuche an DTPA in Modellkla«ranlagen. Abwasserreinigung 37, 707^709. Ternes, T.A., Stumpf, M., Steinbrecher, T., Brenner-WeiM, G. and Haberer, K. (1996) Identi¢cation and detection of new metabolites of DTPA in river water and drinking water. Vom Wasser 87, 275^ 290. GraMho¡, A. and Pottho¡-Karl, B. (1996) Komplexbildner in alkalischen Reinigern. Tens. Surfactant Deterg. 33, 278^288. Schowanek, D., Feijtel, T.C.J., Perkins, C.M., Hartman, F.A., Federle, T.W. and Larson, R.J. (1997) Biodegradation of [S,S], [R,R] and mixed stereoisomers of ethylene diamine disuccinic acid (EDDS), a transition metal chelator. Chemosphere 34, 2375^ 2391. Ro«mheld, V. (1991) The role of phytosiderophores in acquisition of iron and other micronutrients in graminaceous species: An ecological approach. Plant Soil 130, 127^134. Bar-Ness, E., Hadar, Y., Chen, Y., Ro«mheld, V. and Marschner, H. (1992) Short-term e¡ects of rhizosphere microorganisms on Fe uptake from microbial siderophores by maize and oat. Plant Physiol. 100, 451^456. Wire¨n, N., von Ro«mheld, V., Morel, J.L., Guckert, A. and Marschner, H. (1993) In£uence of microorganisms on iron acquisition in maize. Soil Biol. Biochem. 25, 371^376. Wire¨n, N., von Ro«mheld, V., Shioiri, T. and Marschner, H. (1995) Competition between micro-organisms and roots of barley and sorghum for iron accumulated in the root apoplasm. New Phytol. 130, 511^521. Firestone, M.K. and Tiedje, J.M. (1978) Pathway of degradation of nitrilotriacetate by a Pseudomonas species. Appl. Environ. Microbiol. 35, 955^961. Forsberg, C. and Linqvist, G. (1967) Experimental studies on bacterial degradation of NTA. Vatten 23, 265^277. Focht, D. and Joseph, H. (1971) Bacterial degradation of NTA. Can. J. Microbiol. 17, 1553^1556.

FEMSRE 706 29-12-00

104

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

[169] Wong, P.T.S., Liu, D. and McGirr, D.J. (1973) Mechanisms of NTA degradation by a bacterial mutant. Water Res. 7, 1367^1374. [170] Cripps, R.E. and Noble, A.S. (1973) The metabolism of nitrilotriacetate by a Pseudomonad. Biochem. J. 136, 1059^1068. [171] Enfors, S. and Molin, N. (1973) Biodegradation of NTA by bacteria. I: Isolation of bacteria able to grow anaerobically with NTA as sole carbon source. Water Res. 7, 881^888. [172] Liu, D., Wong, P. and Dutka, B. (1973) Studies of a rapid NTA utilizing bacterial mutant. J. Water Pollut. Control Fed. 45, 1728^ 1735. [173] Parks, S. and Stukus, P. (1973) The microbial metabolism of NTA. J. Sci. Lab. Denison Univ. 54, 79^86. [174] Tiedje, J.M., Mason, B.B., Warren, C.B. and Malec, E.J. (1973) Metabolism of nitrilotriacetat by cells of Pseudomonas species. Appl. Microbiol. 25, 811^818. [175] Pickaver, A.H. (1976) The production of N-nitrosoiminodiacetate from NTA and NO3 3 by microorganisms growing in mixed culture. Soil Biol. Biochem. 8, 13^17. [176] Kakii, K., Yamaguchi, H., Iguchi, Y., Teshima, M., Shirakashi, T. and Kuriyama, M. (1986) Isolation and growth characteristics of NTA-degrading bacteria. J. Ferment. Technol. 64, 103^108. [177] Egli, T., Weilenmann, H.U., El-Banna, T. and Auling, G. (1988) Gram-negative, aerobic, nitrilotriacetate-utilizing bacteria from wastewater and soil. Syst. Appl. Microbiol. 10, 297^305. [178] Wanner, U., Kemmler, J., Weilenmann, H.-U., Egli, T., El-Banna, T. and Auling, G. (1990) Isolation and growth of a bacterium able to degrade nitrilotriacetic acid under denitrifying conditions. Biodegradation 1, 31^41. [179] Wehrli, E. and Egli, T. (1988) Morphology of nitrilotriacetate-utilizing bacteria. Appl. Syst. Microbiol. 10, 306^312. [180] Egli, T. and Weilenmann, H.U. (1986) Biodegradation of NTA in the absence of oxygen. Experientia 42, 1061^1062. [181] Auling, G., Busse, H.-J., Egli, T., El-Banna, T. and Stackebrandt, E. (1993) Description of the Gram-negative, obligately aerobic, nitrilotriacetate (NTA)-utilizing bacteria as Chelatobacter heintzii, gen. nov., sp. nov., and Chelatococcus asaccharovorans, gen. nov., sp. nov.. Syst. Appl. Microbiol. 16, 104^112. [182] de Vos, P. and de Ley, J. (1983) Intra- and intergeneric similarities of Pseudomonas and Xanthomonas ribosomal ribonucleic acid cistrons. Int. J. Syst. Bacteriol. 12, 133^142. [183] Ka«mpfer, P., Mu«ller, C., Mau, M., Neef, A., Auling, G., Busse, H.-J., Osborn, A.M. and Stolz, A. (1999) Description of Pseudaminobacter gen. nov. with two new species, Pseudaminobacter salicylatoxidans sp. nov. and Pseudaminobacter de£uvii sp. nov.. Int. J. Syst. Bacteriol. 49, 887^897. [184] Egli, T. and Auling, G. (2001) Genus Chelatobacter, Bergey's Manual of Determinative Bacteriology, accepted. [185] Auling, G. and Egli, T., unpublished information. [186] Egli, T. and Weilenmann, H.-U., unpublished information. [187] Wilberg, E., El-Banna, T., Auling, G. and Egli, T. (1993) Serological studies on nitrilotriacetic acid (NTA)-utilizing bacteria : distribution of Chelatobacter heintzii and Chelatococcus asaccharovorans in sewage treatment plants and aquatic ecosystems. Syst. Appl. Microbiol. 16, 147^152. [188] Bally, M. (1994) Physiology and ecology of nitrilotriacetate degrading bacteria in pure culture, activated sludge and surface waters. Ph.D. Thesis No. 10821, Swiss Federal Institute of Technology, Zu«rich. [189] Meyer, J.-M. and Hohnadel, D. (1992) Use of nitrilotriacetic acid (NTA) by Pseudomonas species through iron metabolism. Appl. Microbiol. Biotechnol. 37, 114^118. [190] Lau¡, J.J., Steele, D.B., Coogan, L.A. and Breitfeller, J.M. (1990) Degradation of the ferric chelate of EDTA by a pure culture of an Agrobacterium sp.. Appl. Environ. Microbiol. 56, 3346^3353. [191] No«rtemann, B. (1992) Total degradation of EDTA by mixed cultures and a bacterial isolate. Appl. Environ. Microbiol. 58, 671^676. [192] Witschel, M., Weilenmann, H.-U. and Egli, T. (1995) Degradation

[193]

[194]

[195] [196]

[197]

[198]

[199] [200] [201]

[202]

[203]

[204]

[205]

[206]

[207]

[208]

[209]

[210]

[211]

[212] [213]

[214]

of EDTA by a bacterial isolate. Poster presented at the 54 Annual Meeting of the Swiss Society for Microbiology, Lugano. Klu«ner, T. (1996) Chemie und Biochemie des mikrobielle EDTAAbbaus. Ph.D. Thesis, Universita«t-Gesamthochschule Paderborn, Paderborn. Henneken, L., No«rtemann, B. and Hempel, D.C. (1995) In£uence of physiological conditions on EDTA degradation. Appl. Microbiol. Biotechnol. 44, 190^197. Wilkinson, S.G. (1970) Cell walls of Pseudomonas species sensitive to ethylenediaminetetraacetic acid. J. Bacteriol. 104, 1035^1044. Ferris, F.G. (1989) Metallic ion interactions with the outer membrane of Gram-negative bacteria. In: Metal Ions and Bacteria (Beveridge, T.J. and Doyle, R.J., Eds.), pp. 295^323. John Wiley and Sons, New York. Jarvis, B.D.W., van Berkum, P., Chen, W.X., Nour, S.M., Fernandez, M.P., Cleyet-Marel, J.C. and Gillis, M. (1997) Transfer of Rhizobium loti, Rhizobium huakuii, Rhizobium ciceri, Rhizobium mediterraneum, and Rhizobium tianshanense to Mesorhizobium gen. nov.. Int. J. Syst. Bacteriol. 47, 895^898. Witschel, M., Nagel, S. and Egli, T. (1997) Identi¢cation and characterization of the two-enzyme system catalyzing the oxidation of EDTA in the EDTA-degrading bacterial strain DSM 9103. J. Bacteriol. 179, 6937^6943. Warren, R.A.J. and Neilands, J.B. (1964) Microbial degradation of the ferrichrome compounds. J. Gen. Microbiol. 35, 459^470. Warren, R.A.J. and Neilands, J.B. (1965) Mechanism of microbial catabolism of ferrichrome A. J. Biol. Chem. 240, 2055^2058. Villavicencio, M. and Neilands, J.B. (1965) An inducible ferrichrome A-degrading peptidase from Pseudomonas FC-1. Biochemistry 4, 1092^1097. Castignetti, D. and Siddiqui, A.S. (1990) The catabolism and heterotrophic nitri¢cation of the siderophore deferrioxyamine B. Biol. Metals 3, 197^203. DeAngelis, R., Forsyth, M. and Castignetti, D. (1993) The nutritional selectivity of a siderophore-catabolizing bacterium. BioMetals 6, 234^238. Harwani, S.C., Roginsky, A., Vallejo, Y. and Castignetti, D. (1997) Further characterization and proposed pathway of deferrioxamine B catabolism. BioMetals 10, 205^213. Zaya, N., Roginsky, A., Williams, J. and Castignetti, D. (1998) Evidence that a deferrioxamine B degrading enzyme is a serine protease. Can. J. Microbiol. 44, 521^527. Winkelmann, G., Schmidtkunz, K. and Rainey, F.A. (1996) characterization of a novel Spirillum-like bacterium that degrades ferrioxamine-type siderophores. BioMetals 9, 78^83. Winkelmann, G., Busch, B., Hartmann, A., Kirchhof, G., Su«ssmuth, R. and Jung, G. (1999) Degradation of desferrioxamines by Azospirillum irakense: Assignment of metabolites by HPLC/electrospray mass spectrometry. BioMetals 12, 255^264. Lau¡, J., Breitfeller, J., Steele, D.B. and Coogan, L. (1993) Degradation of ferric chelates by a pure culture of Agrobacterium sp. US patent 5,252,483. Lau¡, J., Breitfeller, J., Steele, D.B. and Coogan, L. (1994) Pure culture of Agrobacterium sp. which degrades ferric chelates. US patent 5,364,786. Henneken, L., No«rtemann, B. and Hempel, D.C. (1998) Biological degradation of EDTA: Reaction kinetics and technical approach. J. Chem. Technol. Biotechnol. 73, 144^152. Henneken, L., Klu«ner, T., No«rtemann, B. and Hempel, D.C. (1994) Abbau von EDTA mit freien und immobilisierten Bakterien. gwf Wasser Abwasser 135, 354^358. Bru«ggenthies, A. (1996) Biologische Reinigung EDTA-haltiger Abwa«sser. FIT-Verlag, Paderborn. Henneken, L., Bru«ggenthies, A., No«rtemann, B. and Hempel, D.C. (1996) Teilstrombehandlung EDTA-haltiger Abwa«sser mittels Bio¢lm-Wirbelbettreaktoren. Chem. Ing. Tech. 68, 310^314. ë kologie Nitrilotriacetat Wilberg, E. (1989) Zur Physiologie und O

FEMSRE 706 29-12-00

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

[215] [216]

[217]

[218]

[219]

[220]

[221]

[222]

[223]

[224]

[225]

[226]

[227] [228]

[229]

[230]

[231]

[232]

[233]

(NTA) abbauender Bakterien. Ph.D. Thesis No. 9015, Swiss Federal Institute of Technology, Zu«rich. Cripps, R.E. and Noble, A.S. (1972) The microbial metabolism of nitrilotriacetate. Biochem. J. 130, 31P^32P. Uetz, T., Schneider, R., Snozzi, M. and Egli, T. (1992) Puri¢cation and characterization of a two-component monooxygenase that hydroxylates nitrilotriacetate from `Chelatobacter' strain ATCC 29600. J. Bacteriol. 174, 1179^1188. Uetz, T. and Egli, T. (1993) Characterization of an inducible membrane-bound iminodiacetate dehydrogenase from Chelatobacter heintzii ATCC 29600. Biodegradation 3, 423^434. Uetz, T.A. (1992) Biochemistry of nitrilotriacetate degradation in obligately aerobic, Gram-negative bacteria. Ph.D. Thesis No. 9722, Swiss Federal Institute of Technology, Zu«rich. Knobel, H.-R., Egli, T. and van der Meer, J.R. (1996) Cloning and characterization of the genes encoding nitrilotriacetate monooxygenase of Chelatobacter heintzii ATCC 29600. J. Bacteriol. 178, 6123^ 6132. Xu, Y., Mortimer, M.W., Fisher, T.S., Kahn, M.L., Brockman, F.J. and Xun, L. (1997) Cloning, sequencing, and analysis of a gene cluster from Chelatobacter heintzii ATCC 29600 encoding nitrilotriacetate monooxygenase and NADH:Flavin mononucleotide oxidoreductase. J. Bacteriol. 179, 1112^1116. Xi, L., Squires, C.H., Monticello, D.J. and Childs, J.D. (1997) A £avin reductase stimulates DszA and DszC proteins of Rhodococcus erythropolis IGTS8 in vitro. Biochem. Biophys. Res. Commun. 230, 73^75. Thibaut, D., Ratet, N., Bisch, D., Faucher, D., Debussche, L. and Blanche, F. (1995) Puri¢cation of the two-enzyme system catalyzing the oxidation of the D-proline residue of pristinamycin IIB during the last step of pristinamycin IIA biosynthesis. J. Bacteriol. 177, 5199^5205. Kendrew, S.G., Harding, S.E., Hopwood, D.A. and Marsh, E.N.G. (1995) Identi¢cation of a £avin:NADH oxidoreductase involved in the biosynthesis of actinorhodin. J. Biol. Chem. 270, 17339^17343. Parry, R.J. and Li, W. (1997) An NADPH :FAD oxidoreductase from the valanimycin producer, Streptomyces viridifaciens. J. Biol. Chem. 272, 23303^23311. Knobel, H.-R. (1997) Genetic study of bacterial nitrilotriacetate degrading enzymes. Ph.D. Thesis No. 12146, Swiss Federal Institute of Technology, Zu«rich. Kertesz, M.A., Schmidt-Larbig, K. and Wuest, T. (1999) A novel reduced £avin mononucleotide-dependent methanesulfonate sulfonatase encoded by the sulfur-regulated msu operon of Pseudomonas aeruginosa. J. Bacteriol. 181, 1464^1473. Fontecave, M., Coves, J. and Pierre, J.-L. (1994) Ferric reductases or £avin reductases? BioMetals 7, 3^8. Jenal-Wanner, U. and Egli, T. (1993) Anaerobic degradation of nitrilotriacetate (NTA) in a denitrifying bacterium: puri¢cation and characterization of the NTA dehydrogenase. Appl. Environ. Microbiol. 59, 3350^3359. Kemmler, J. (1992) Biochemistry of nitrilotriacetate degradation in the facultativly denitrifying bacterium TE11. Ph.D. Thesis No. 9983, Swiss Federal Institute of Technology, Zu«rich. Steenkamp, D.J. and Gallup, M. (1978) The natural £avoprotein electron acceptor of trimethylamine dehydrogenase. J. Biol. Chem. 253, 4086^4089. Witschel, M., Egli, T., Zehnder, A.J.B., Wehrli, E. and Spycher, M. (1999) Transport of EDTA into cells of the EDTA-degrading strain DSM 9103. Microbiology 154, 973^983. Klu«ner, T., Hempel, D.C. and No«rtemann, B. (1998) Metabolism of EDTA and its metal chelates by whole cells and cell-free extracts of strain BNC1. Appl. Environ. Microbiol. 49, 194^201. Witschel, M. (1999) Biochemical and physiological characterisation of a bacterial isolate able to grow with EDTA and other aminopolycarboxylic acids. Ph.D. Thesis No. 12967, Swiss Federal Institute of Technology, Zu«rich.

105

[234] Payne, J.W., Bolton, H.J., Campbell, J.A. and Xun, L. (1998) Puri¢cation and characterization of EDTA monooxygenase from the EDTA-degrading bacterium BNC1. J. Bacteriol. 180, 3823^ 3827. [235] Witschel, M. and Egli, T. (1998) Puri¢cation and characterization of a lyase from the EDTA-degrading bacterial strain DSM 9103 that catalyzes the splitting of [S,S]-ethylenediaminedisuccinate, a structural isomer of EDTA. Biodegradation 8, 419^428. [236] Bally, M., Wilberg, E., Ku«hni, M. and Egli, T. (1994) Growth and regulation of enzyme synthesis in the nitrilotriacetic acid (NTA)degrading Chelatobacter heintzii ATCC 29600. Microbiology 140, 1927^1936. [237] Bally, M. and Egli, T. (1996) Dynamics of substrate consumption and enzyme synthesis in Chelatobacter heintzii during growth in carbon-limited chemostat culture with di¡erent mixtures of glucose and nitrilotriacetate (NTA). Appl. Environ. Microbiol. 62, 133^140. [238] Egli, T., Ka«ppeli, O. and Fiechter, A. (1982) Mixed substrate growth of methylotrophic yeasts in chemostat culture: in£uence of the dilution rate on the utilisation of a mixture of glucose and methanol. Arch. Microbiol. 131, 8^13. [239] Tempest, D.W., Neijssel, O.M. and Zevenboom, W. (1983) Properties and performance of microorgansisms in laboratory culture ; their relevance to growth in natural ecosystems. S. Soc. Gen. Microbiol. 34, 119^152. [240] Egli, T. (1995) The ecological and physiological signi¢cance of the growth of heterotrophic microorganisms with mixtures of substrates. Adv. Microb. Ecol. 14, 305^386. [241] Barford, J.P., Pamment, N.B. and Hall, R.J. (1982) Lag phases and transients. In: Microbial Population Dynamics (Bazin, J., Ed.), pp. 55^89. CRC Press, Boca Raton, FL. [242] Neef, A. (1997) Anwendung der in situ-Einzelzell-Identi¢zierung von Bakterien zur Populationsanalyse in komplexen mikrobiellen Biozo«nosen. Ph.D. Thesis, Technische Universita«t Mu«nchen, Mu«nchen. [243] Bjo«rndal, H., Bouveng, H.O., Solyom, P. and Werner, J. (1972) Biochemical stability of some metal chelates. Vatten 28, 5^16. [244] Swisher, R.D., Taulli, T.A. and Malec, E.J. (1973) Biodegradation of NTA metal chelates in river water. In: Trace Metals and Metal^ Organic Interactions in Natural Waters (Singer, P.C., Ed.). Ann Arbor Sc. Publ., Inc., Ann Arbor, MI. [245] Walker, A.P. (1975) Ultimate biodegradation of nitrilotriacetate in the presence of heavy metals. Prog. Water Technol. 7, 555^560. [246] Gundernatsch, H. (1975) Biologischer Abbau von Schwermetallkomplexen der Nitrilotriessigsa«ure in Laborbelebtschlammanlagen. gwf Wasser Abwasser 116, 512^517. [247] Firestone, M.K. and Tiedje, J.M. (1975) Biodegradation of metal^ nitrilotriacetate complexes by a Pseudomonas species: mechanism of reaction. Appl. Microbiol. 29, 758^764. [248] Bolton, H.J., Girvin, D.C., Plymale, A.E., Harvey, S.D. and Workman, D.J. (1996) Degradation of metal^nitrilotriacetate complexes by Chelatobacter heintzii. Environ. Sci. Technol. 30, 931^938. [249] Palumbo, A.V., Lee, S.Y. and Borman, P. (1994) The e¡ect of media composition on EDTA degradation by Agrobacterium sp.. Appl. Biochem. Biotechnol. 45/46, 811^822. [250] Satroutdinov, A.D., Dedyukhina, E.G., Chistyakova, T.I., Witschel, M., Minkevich, I.G., Eroshin, V.K. and Egli, T. (2000) Degradation of metal^EDTA complexes by resting cells of the bacterial strain DSM 9103. Environ. Sci. Technol. 34, 1715^1720. [251] Xun, L., Reeder, R.B., Plymale, A.E., Girvin, D.C. and Bolton, H. (1996) Degradation of metal nitrilotriacetate complexes by nitrilotriacetate monooxygenase. Environ. Sci. Technol. 30, 1752^ 1755. [252] Jenal-Wanner, U. (1991) Anaerobic degradation of nitrilotriacetate in a denitrifying bacterium : puri¢cation and characterization of the nitrilotriacetate dehydrogenase/nitrate reductase complex. Ph.D. Thesis No. 9531, Swiss Federal Institute of Technology, Zu«rich. [253] Silver, S. and Lusk, J.E. (1987) Bacterial magnesium manganese and

FEMSRE 706 29-12-00

106

M. Bucheli-Witschel, T. Egli / FEMS Microbiology Reviews 25 (2001) 69^106

zinc transport. In: Ion Transport in Prokaryotes (Rosen, B.P. and Silver, S., Ed.). Academic Press, Inc., San Diego, CA. [254] Martell, A.E. and Smith, R.M. (1974) Critical Stability Constants, Vol. 1: Amino Acids. Plenum Press, New York. [255] Smith, G.S. and Hoard, J.L. (1959) The structure of dihydrogen

ethylenediaminetetraacetatoaquonickel(II). J. Am. Chem. Soc. 81, 556. [256] Lind, M.D., Hamor, M.J., Hamor, T.A. and Hoard, J.L. (1964) Stereochemistry of EDTA complexes. II. The structure of crystalline Rb[Fe(OH2 )Y]H2 O. Inorg. Chem. 3, 34^43.

FEMSRE 706 29-12-00

Lihat lebih banyak...

Comentarios

Copyright © 2017 DATOSPDF Inc.