Environmental chemistry of phosphonates

Water Research 37 (2003) 2533–2546

Review

Environmental chemistry of phosphonates Bernd Nowack* . Swiss Federal Institute of Technology Zurich Institute of Terrestrial Ecology (ITO), (ETH), Grabenstrasse 3, . CH-Schlieren 8952, Switzerland Received 16 August 2002; received in revised form 21 January 2003; accepted 31 January 2003

Abstract Phosphonates are anthropogenic complexing agents containing one or more C–PO(OH)2 groups. They are used in numerous technical and industrial applications as chelating agents and scale inhibitors. Phosphonates have properties that differentiate them from other chelating agents and that greatly affect their environmental behavior. Phosphonates have a very strong interaction with surfaces, which results in a significant removal in technical and natural systems. Due to this strong adsorption, little or no remobilization of metals is expected. No biodegradation of phosphonates during water treatment is observed but photodegradation of the Fe(III)-complexes is rapid. Aminopolyphosphonates are also rapidly oxidized in the presence of Mn(II) and oxygen and stable breakdown products are formed that have been detected in wastewater. The lack of information about phosphonates in the environment is linked to analytical problems of their determination at trace concentrations in natural waters. Further method development is urgently needed in this area, including speciation of these compounds. With the current knowledge on speciation, we can conclude that phosphonates are mainly present as Ca and Mg-complexes in natural waters and therefore do not affect metal speciation or transport. r 2003 Elsevier Science Ltd. All rights reserved. Keywords: Phosphonates; Chelating agents; Adsorption; Heavy metals; Degradation; Speciation

Contents 1.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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2.

Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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3.

Analysis of phosphonates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Analytical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Concentrations in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . .

2536 2536 2537

4.

Surface reactions . . . . . . . . . . . . . . . . . 4.1. Adsorption . . . . . . . . . . . . . . . . . 4.2. Dissolution of minerals . . . . . . . . . . . 4.3. Remobilization of metals . . . . . . . . . . 4.4. Precipitation . . . . . . . . . . . . . . . . 4.5. Inhibition of dissolution and precipitation .

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*Tel.: +41-1-633-61-60; fax: +41-1-633-11-23. E-mail address: [email protected] (B. Nowack). 0043-1354/03/$ - see front matter r 2003 Elsevier Science Ltd. All rights reserved. doi:10.1016/S0043-1354(03)00079-4

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B. Nowack / Water Research 37 (2003) 2533–2546

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5.

Degradation . . . . . . . . . . . . . . . . . 5.1. Biodegradation . . . . . . . . . . . . . 5.2. Photodegradation . . . . . . . . . . . . 5.3. Chemical degradation . . . . . . . . . 5.4. Degradation during oxidation processes

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2539 2539 2540 2540 2541

6.

Speciation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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7.

Behavior during wastewater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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8.

Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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

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

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1. Introduction Phosphonic acids, compounds containing the Lewis acid moiety R-CP(O)(OH)2, are characterized by a stable, covalent carbon to phosphorous bond. The corresponding anions of the phosphonic acids are called phosphonates. The most commonly used phosphonates are structural analogues to the well-known aminopolycarboxylates such as ethylenediaminetetra acetate (EDTA) and nitrilotriacetate (NTA). The environmental fate of these aminopolycarboxylate chelating agents has received considerable attention [1–5]. Much less is known about the fate and behavior of the corresponding phosphonates in the environment [4,6,7]. The existing reviews are either several years old and therefore do not cover the newest literature [6] or focus on toxicology and risk assessment based on the limited data that were available at that time [7]. What is missing is an overview of the chemistry of these compounds which can help us to understand and predict the environmental behavior of these compounds more accurately and that can be the basis for a refined risk assessment. The aim of this review is therefore to provide an overview of the current knowledge of the environmental chemistry of phosphonates. It concentrates on polyphosphonates, compounds containing more than one phosphonic acid group, and especially aminopolyphosphonates, compounds containing several phosphonate and one or more amine groups. Glyphosate, a herbicide containing a phosphonate, a carboxylate and an amine functional group, is not discussed in detail in this review. There is, however, much information available about the environmental chemistry and behavior of this compound [8–10]. This review starts with a short description of the properties of phosphonates and their analysis. Phosphonates have a very strong interaction with surfaces and the section discussing the surface reaction follows: adsorption, dissolution of minerals, remobilization of metals, precipitation of phosphonates and inhibition of precipitation of minerals are covered. In the degradation

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section biodegradation, photodegradation, chemical degradation and degradation during oxidation processes are discussed. The speciation of phosphonates in the environment covers the next section, which is followed by a discussion of their environmental behavior. This section contains a summary of the data on measured concentrations of phosphonates and their behavior during wastewater treatment.

2. Properties Table 1 lists the abbreviations, names and structures of the phosphonates discussed in this review. These compounds are known under many different abbreviations that vary between the disciplines and countries and have changed with time. Phosphonates are effective chelating agents according to the IUPAC definition that chelation involves coordination of more than one sigmaelectron pair donor group from the same ligand to the same central atom. Phosphonates are used as chelating agents in many applications, e.g. in pulp, paper and textile industry to complex heavy metals in chlorine-free bleaching solutions that could inactivate the peroxide. In medicine phosphonates are used to chelate radionuclides for bone cancer treatments [11]. A recent IUPAC Technical Report [12] critically evaluates the available experimental data on stability constants of proton and metal complexes for phosphonic acids. It presents high-quality data as ‘‘recommended’’ or ‘‘provisional’’ constants while for example, all constants for DTPMP have been rejected due to insufficient purity of the parent compound. This report will be of great use for all future speciation calculations and should be the sole source of stability constants when ever possible. The stability of the metal complexes increases with increasing number of phosphonic acid groups. Fig. 1 shows that the monophosphonate aminomethylphosphonic acid (AMPA) has the lowest stability constants

Other abbreviations also in use

HEDPA, HEBP

ATMP, NTP, NTPH, NTPO

EDTP, EDTPH, ENTMP, EDTMPO, EDTMPA

DETPMP, DTPPH, DETPMPA, DETPMPO

PBTCA

Abbreviation

HEDP

NTMP

EDTMP

DTPMP

PBTC

Phosphonobutane-tricarboxylic acid

Diethylenetriaminepentakis (methylenephosphonic acid)

1,2-Diaminoethanetetrakis (methylenephosphonic acid)

Nitrilotris(methylenephosphonic acid)

1-Hydroxyethane(1,1-diylbisphosphonic acid)

Name

Table 1 Abbreviations, names, and structures of the phosphonates covered in this review

PO (OH )2 OH

(HO) 2OP

HOOC

(HO)2OP

(HO)2OP

(HO) 2OP

(HO)2OP

N

N

N

COOH

C

PO(OH)2

PO (OH )2

C

(HO) 2OP

H3 C

Structure

PO(OH)2

N

PO(OH)2

PO (OH)2

COOH

N

N

PO(OH) 2

PO(OH)2

PO(OH)2

B. Nowack / Water Research 37 (2003) 2533–2546 2535

B. Nowack / Water Research 37 (2003) 2533–2546

2536 20

1 10-6

15

8 10-7

EDTMP

10

5

HEDP

NTMP IDMP AMPA

concentration (M)

log K (M + HL)

3-

H EDTMP

ZnHEDTMP

5

2-

H EDTMP

5-

6

6 10-7 H EDTMP

44-

4

4 10-7

ZnH EDTMP 2

2 10-7

3-

ZnH EDTMP

-

3

H EDTMP

0

7-

HEDTMP

7

2+

Mn

2+

Fe

2+

Co

2+

Ni

2+

Cu

2+

Zn

Fig. 1. Stability constants of 1:1 complexes (M+HL) with transition metals of AMPA, IDMP, HEDP, NTMP and EDTMP (Irving–Williams series) with data from [12].

and EDTMP with 4 phosphonic aid groups the highest. The log K values of the different transition metal complexes follow the Irving–Williams series Mn2+o Fe2+oCo2+oNi2+oCu2+>Zn2+. Fig. 2 shows a speciation diagram for the system ZnEDTMP calculated with the constants from [12] with the speciation program ChemEQL [13]. This calculation shows that in the pH range found in technical applications and in natural waters a large number of possible complexes with different degree of protonation and charge exist. At pH 6 the species H4EDTMP4
, ZnH3EDTMP3
, ZnH2EDTMP4
and ZnHEDTMP5
occur at a percentage of more than 5% of total EDTMP. Complexation of other metals by other phosphonates is similar and at each pH value several species coexist. Phosphonates are not only chelating agents but also very potent inhibitors of mineral precipitation and growth. This effect works at concentrations well below the amount needed to chelate all metals. An important industrial use of phosphonates is in cooling waters, desalination systems and in oil fields to inhibit scale formation, e.g. barium sulfate or calcium carbonate precipitation. Phosphonates are also used in medicine to treat various bone and calcium metabolism diseases [14]. In detergents phosphonates are used as a combination of chelating agent, scale inhibitor and bleach stabilizer [15]. Phosphonates are highly water-soluble while the phosphonic acids are only sparingly soluble. Phosphonates are not volatile and poorly soluble in organic solvents. More detailed data on the physicochemical properties of the phosphonates can be found in reference [7]. The consumption of phosphonates was 56,000 tons worldwide in 1998 [16] and 16,000 tons in Europe in 1999 [4]. Data about the distribution among the various phosphonates are available for Europe and the US [6], for the Netherlands [7] and for Germany [4]. HEDP and DTPMP are the most important phosphonates based on the used volumes.

0 2

4

6

8

10

12

pH Fig. 2. Speciation of 1 mM EDTMP in the presence of 1 mM Zn. The diagram has been calculated using the constants from [12].

The toxicity of phosphonates to aquatic organisms is low [6,7,17]. Reported values for 48 h LC50 values for fish are between 0.1 and 1.1 mM [18,19]. Also the bioconcentration factor for fish is very low [20,21]. Phosphonates are poorly absorbed in the gastrointestinal tract and most of the absorbed dose was rapidly excreted by the kidneys [22]. Human toxicity is also low which can be seen in the fact that phosphonates are used to treat various diseases [14,23].

3. Analysis of phosphonates 3.1. Analytical methods The absence of a reliable trace analytical method for phosphonates results in a lack of detailed information about the environmental behavior of phosphonates. Most of the current methods for phosphonate determination have detection limits above the expected natural concentrations or suffer from interferences in natural samples. The standard method for the determination of phosphonates is ion-chromatography followed by postcolumn reaction with Fe(III) and detection of the Fe(III)-complexes at 300–330 nm [24–26]. This method has a detection limit of about 2–10 mM. Other methods have been developed based on post-column oxidation of the phosphonate to phosphate and detection of phosphate with the molybdenum blue method [27]. Ionchromatography with pulsed amperometric detection of amine-containing phosphonates [28], ion-chromatography with indirect photometric detection [29] and capillary electrophoresis with indirect photometric detection have also been described [30]. These methods all have high detection limits of 1 mM or more and are therefore not suitable for natural systems. A very powerful method is the derivatization of the phosphonic acid group with diazomethane and

B. Nowack / Water Research 37 (2003) 2533–2546

separation and detection of the derivatives by HPLCMS [31]. This method is however, not applicable to natural waters due to interference by the major cations and anions of the water matrix. The only method with a low enough detection limit in natural samples is an ionpair HPLC method with precolumn formation of the Fe(III)-complexes [32]. The phosphonates can be measured with a detection limit of 0.05 mM in natural waters and wastewaters. The method, however, is not able to quantify bisphosphonic acids such as HEDP at low concentrations. This is a major drawback because HEDP is one of the most used phosphonates [4,6]. The breakdown products of the Mn(II)-catalyzed degradation of NTMP [33], iminodimethylenephosphonic acid (IDMP) and N-formyl-iminodimethylenephosphonic acid (FIDMP), can be detected after derivatization of the aldehyde group in FIDMP by 2,4-dinitrophenylhydrazine and derivatization of the imine-group in IDMP by 9-fluorenyl methylchloroformate [34]. A detection limit of 0.01 mM FIDMP and 0.02 mM IDMP has been achieved. Anion-exchange chromatography coupled to ICP-MS is able is a very promising method for chelating agent analysis [35,36]. The method is also applicable to phosphonates and it has been shown that CuEDTMP can be determined with a very low detection limit in the nanomolar range. Preconcentration of phosphonates from natural water samples using different adsorbents has been tested [37]. It was found that the investigated phosphonates HEDP, NTMP, and EDTMP differed so much in their chemical behavior that a simultaneous enrichment from natural samples cannot be achieved. Successful preconcentration of the phosphonates NTMP, EDTMP and DTPMP from natural waters or wastewaters was achieved using freshly precipitated CaCO3 [32]. Recoveries at the 1 mM level were 95–102% for an influent sample of a wastewater treatment plant. 3.2. Concentrations in the environment No measurements of phosphonates in natural samples have been reported and only data for wastewaters are available. This is mainly due to the fact that most analytical methods are not able to quantify phosphonates in natural waters at low concentrations. Phosphonates have been measured in Swiss wastewater treatment plants (WWTP) [38]. The concentrations of NTMP were between o0.05 and 0.85 mM, of EDTMP between o0.05 and 0.15 mM and of DTPMP between o0.05 and 1.7 mM. The highest concentration of DTPMP was found in a WWTP influenced by textile industry. Effluent samples from all investigated WWTP were with the exception of one case always below the detection limit. Another WWTP influenced by textile industry contained NTMP concentrations in the influent between 0.2 and 1.1 mM [39].

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The oxidative breakdown products of NTMP, IDMP and FIDMP, have been detected in two WWTPs receiving water from textile industry at concentrations of 0.08 and 0.015 mM FIDMP and 0.49 and 0.3 mM IDMP in the influent [34]. The expected concentrations in rivers are maximal 0.1 mM and with adsorption/photodegradation included about 1–4 nM for NTMP and 25 nM for HEDP [6,7]. Locally higher concentrations can be expected because of intermitted discharge of cooling tower water.

4. Surface reactions 4.1. Adsorption Phosphonates adsorb very strongly onto almost all mineral surfaces. This behavior distinguishes them from the corresponding aminocarboxylates, which exhibit much weaker interaction with mineral surfaces, especially near neutral pH [40]. Some of the investigated adsorbents for phosphonates are calcite [41], clays [42,43], aluminum oxides [44–46], iron oxides [47–49], zinc oxide [49], hydroxyapatite [50,51] and barite [52]. For all those compounds very strong adsorption is observed in the pH range of natural waters. Natural materials are also very potent adsorbents for phosphonates, for example sewage sludge [20,21,39,53,54], sediments [54] and soils [55]. Most of these studies, however, have not considered that metal ions might significantly alter the adsorption of a chelating agent [56]. However, no influence of Fe(III), Zn, and Cu(II) on phosphonate adsorption onto goethite was observed [49]. This was explained by the very strong adsorption of the uncomplexed phosphonate, which resulted in a dissociation of the complex at the surface and separate adsorption of the metal and the phosphonate onto different surface sites. Fig. 3 shows the adsorption of NTMP and the NTMP complexes with Zn, Cu and Fe(III). Complete adsorption is observed up to a pH of 8 and no influence of the complexed metal on the shape of the adsorption edge can be seen. In the pH range of natural waters adsorption is therefore very strong. Other phosphonates, e.g. HEDP, EDTMP and DTPMP, adsorb in a similar manner to NTMP. Ca has a very strong positive effect on phosphonate adsorption [49]. In the presence of mM Ca concentrations, phosphonates were completely adsorbed up to pH of 12. The maximum surface concentration of phosphonates was also greatly enhanced in the presence of Ca. This effect can be explained by the formation of ternary surface-phosphonate-Ca complexes. Precipitation of Caphosphonates on the surface can be ruled out [49]. When evaluating the adsorptive capacity of a surface towards phosphonates in a natural system, it is therefore

B. Nowack / Water Research 37 (2003) 2533–2546

100

1 10-6

80

8 10-7

Fe(III)NTMP

% adsorbed

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60 40

NTMP ZnNTMP CuNTMP Fe(III)NTMP NTMP/ 1 mM Ca

20

NTMP 2 mM Ca 1 µM Zn 1 µM Cu

6 10-7 4 10-7 2 10-7

0

0

4

6

8

10

12

4

4.5

5

5.5

pH Fig. 3. Adsorption of 10 mM NTMP onto goethite in the absence and presence of equimolar Zn, Cu, and Fe(III) and 1 mM Ca. Reprinted with permission from [49]. Copyright (1999) American Chemical Society.

necessary to conduct the adsorption experiments under natural Ca concentrations. 4.2. Dissolution of minerals Dissolution of a mineral phase by chelating agents can be explained in terms of a ligand exchange process and is related to the concentration of surface bound ligands. The ligands weaken the metal–oxygen bonds on the surface and enhance the release of metal ions from the surface into the adjacent solution [57]. Reactions with iron oxides are especially of great importance regarding the speciation of the ligand in solution due to the very strong Fe(III)-complexes. Reactions like this have been observed in subsurface systems and have a pronounced influence on the mobility of heavy metals [58,59]. Very little is known about the dissolution of iron oxides by phosphonates. It was observed that HEDP significantly mobilized Fe from natural sediments but no information was given about the pH value of the experiments [60]. No enhanced solubilization of Fe from river sediment was observed at pH 3 by 0.01 M NTMP [61]. The concentration of the Fe(III)-complex in the presence of an iron oxide phase can be calculated when the stability constants of the Fe(III)-complexes are known. Fig. 4 shows the calculated Fe(III)NTMP concentration in a system with NTMP and hydrous ferric oxide (HFO). The speciation has been calculated with the published stability constants for metal-NTMP complexes [12] and the NTMP-Fe(III) stability constants from [62] using the program ChemEQL [13]. The formation of Fe(III)NTMP is important at pH values below 6 in the absence of other metal ions. At pH above 7 NTMP is present as uncomplexed ligand. 1 mM Ca and equimolar Zn depress the formation of

6

6.5

7

7.5

8

pH Fig. 4. Speciation of 1 mM NTMP in the presence of HFO and in the absence and presence of 1 mM Zn and Cu and 2 mM Ca without considering adsorption of the phosphonate. Log K values from [12] and [62].

Fe(III)NTMP slightly. Cu which forms the strongest complexes with NTMP has the largest influence on Fe(III)NTMP formation. We can therefore conclude that dissolution reactions are able to occur at low pH. However, the strong adsorption of phosphonates, especially at low pH, will limit the formation of dissolved Fe(III)NTMP complexes and therefore no dissolution will occur at low, environmentally relevant concentrations. 4.3. Remobilization of metals Metals adsorbed onto a mineral surface can be solubilized by chelating agents. This process has always been mentioned as one of the most adverse effects of elevated chelating agent concentrations in the environment [63]. Metal adsorption in the presence of phosphonates has been studied. There is an increase in Cu adsorption in the presence of phosphonates at low pH, which is caused by electrostatic effects [49]. At high pH there is a mobilization of Cu due to the formation of dissolved Cu-phosphonate complexes. Fig. 5 shows as an example the influence of EDTMP on Cu adsorption onto goethite. Overall, the influence of phosphonates on metal adsorption in the natural pH range from 4 to 8 is weak. We can therefore expect that phosphonates have only a slight influence on metal remobilization in natural systems. This was actually found during the study of metal mobilization from river sediments by the phosphonate HEDP [60]. The only metal to be remobilized was Fe whereas Zn, Cr, Ni, Cu, Pb and Cd were not increased compared to a blank sample and only dissolution of iron oxides was observed. Remobilization of Cu, Cd, and Pb from river sediment was only observed at NTMP concentrations above 0.1 mM [61].

B. Nowack / Water Research 37 (2003) 2533–2546

% Cu adsorbed

100 80 60 40 20

Cu alone Cu with NTMP

0 4

6

8

10

12

pH Fig. 5. Adsorption of 10 mM Cu onto goethite in the absence and presence of 10 mM NTMP. Reprinted with permission from [49]. Copyright (1999) American Chemical Society.

We can therefore conclude that phosphonates probably have only a marginal influence on metal mobilization in the environment. 4.4. Precipitation In many applications, phosphonates are added to waters containing high concentrations of dissolved ions to prevent the formation of precipitates. However, due to the insolubility of some metal-phosphonates, the phosphonates itself can precipitate. This phenomenon often occurs in oil field applications when phosphonates are injected into the subsurface and are left to interact with calcium-containing formation waters [64,65]. The solubility of precipitates of NTMP with divalent metals increases in the order CaoBaoSroMg [66]. The insoluble Ca precipitates of DTPMP [67,68], NTMP [69], and HEDP [70] and the precipitates of NTMP with Fe(II) [71] and Fe(III) [72] have been investigated in detail. Insoluble products of HEDP are also formed with heavy metals such as Pb and Cd [73]. The precipitates are important in oil field applications or in technical systems where high phosphonate and high ion concentration occur simultaneously. In natural waters or wastewaters, the phosphonate or Ca concentrations are far too low to exert any influence on phosphonate concentrations. The solubility of NTMP in the presence of 1 and 5 mM Ca is always above 200 mM [49]. In natural waters precipitation reactions are therefore not important. 4.5. Inhibition of dissolution and precipitation Scale formation, e.g. precipitation of calcium carbonate or calcium sulfate, is a significant problem in commercial water treatment processes including cooling water technology, desalination and oil field applications. This scale formation can be alleviated by the use of

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chemical water treatment additives, known as ‘‘threshold inhibitors’’. Phosphonic acids are among the most potent scale inhibitors next to the polyphosphates. They poison the crystal growth at concentrations far below stoichiometric amounts of the reactive cations. Models for this poisoning include inhibition of nucleation, adsorption onto growth sites, distortion of the crystal lattice, changes in surface charge and association with precursors of crystal formation [74,75]. The morphology of crystals formed in the presence of phosphonates is markedly different from those in the absence of phosphonates [76]. Phosphonates limit the size of the growing crystals and produce a lag phase in which crystal growth is greatly reduced [77]. It was found that the ability of different phosphonates to inhibit crystal growth can be interpreted in terms of the Langmuir adsorption model with the strongest inhibitory effect from compounds that adsorb most strongly [78]. Due to their inhibitory effect on crystal growth it has been argued that phosphonates may have an adverse effect on phosphate elimination by precipitation with iron or aluminum salts during wastewater treatment [79,80]. It was found that the phosphonates had an influence on flocculation but it was possible to compensate for it by increased addition of flocculating agent. The resulting particulate precipitation products were stabilized by the dispersing action of the phosphonates and not retained in the sand filter. Another study, however, found no influence of HEDP on phosphate elimination [60].

5. Degradation 5.1. Biodegradation Phosphonates are similar to phosphates except that they have a carbon–phosphorous (C–P) bond in place of the carbon–oxygen–phosphorous (C–O–P) linkage. Due to their structural similarity to phosphate esters, phosphonates often act as inhibitors of enzymes due in part to the high stability of the C–P bond [81]. In nature bacteria play a major role in phosphonate biodegradation. The first phosphonate to be identified to occur naturally was 2-aminoethylphosphonic acid [82]. It is found in plants and many animals, mostly in membranes. Phosphonates are quite common among different organisms, from prokaryotes to eubacteria and fungi, mollusks, insects and others but the biological role of the natural phosphonates is still poorly understood [83]. Due to the presence of natural phosphonates in the environment, bacteria have evolved the ability to metabolize phosphonates as nutrient sources. Those bacteria able of cleaving the C–P bond are able to use phosphonates as a phosphorous source for growth.

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Aminophosphonates can also be used as sole nitrogen source by some bacteria [84]. The polyphosphonate chelating agents discussed here differ greatly from natural phosphonates such as 2aminoethylphosphonic acid, because they are much larger, carry a high negative charge and are complexed with metals. Biodegradation tests with sludge from municipal sewage treatment plants with HEDP and NTMP showed no indication for any degradation based on CO2 formation [18,20,21]. An investigation of HEDP, NTMP, EDTMP and DTPMP in standard biodegradation tests also failed to identify any biodegradation [53]. It was noted, however, that in some tests due to the high sludge to phosphonate ratio, removal of the test substance from solution observed as loss of DOC was observed. This was attributed to adsorption rather than biodegradation because no accompanying increase in CO2 was observed. However, bacterial strains capable of degrading aminopolyphosphonates and HEDP under P-limited conditions have been isolated from soils, lakes, wastewater, activated sludge and compost [85]. The phosphonate phosphonobutane-tricarboxylic acid (PBTC) was also rapidly degraded by microbial enrichment cultures from a variety of ecosystems under conditions of low phosphate availability [86]. The effects of other more accessible P sources on phosphonate uptake and degradation are of great environmental importance. Many environments such as activated sludge, sediments and soils that act as a sink for phosphonates are not characterized by a lack of P most of the time. Because phosphonates are utilized almost exclusively as P-source, little biodegradation can be expected under these conditions. It has been demonstrated, however, that simultaneous phosphate and phosphonate utilization by bacteria can occur [87]. Adsorption of chelating agents by surfaces has been shown to decrease the biodegradability. The easily biodegradable NTA for example is much slower degraded when adsorbed to mineral surfaces [88]. It can be expected that phosphonates with their higher affinity to surfaces are much slower degraded in a heterogeneous compared to a homogeneous system. This was found to be the case for N-phosphonomethylglycine, the phosphonate-containing herbicide glyphosate [89]. Phosphonates are therefore similar to EDTA [3,90] in that little or no biodegradation is observed in natural systems but that microorganisms have been isolated from these environments capable of degrading the compound. 5.2. Photodegradation Photodegradation of the Fe(III)-complexes is an important pathway of aminopolycarboxylate elimination

in the environment [91]. Phosphonates have a similar reactivity. In distilled water and in the presence of Ca no photodegradation of HEDP was observed but the addition of Fe(III) and Cu(II) resulted in rapid photodegradation [20,92]. The mechanism of Fe(III)EDTMP photodegradation [93] is equivalent to the photodegradation of Fe(III)EDTA [94]. Fe(III)EDTMP is degraded in a stepwise process from the parent compound through ethylenediaminetrimethylenephosphonate and ethylenediaminedimethylenephosphonate to ethylenediaminemonomethylenephosphonate which is stable in the presence of Fe(III) and light. For EDTA the photodegradation of the Fe(III)-complexes is the major elimination pathway in natural waters [91]. We can therefore expect that photodegradation is also very important for the fate of dissolved phosphonates in surface waters. The photodegradation products of Fe(III)EDTA are readily biodegradable, but this is not the case for phosphonates [95]. 5.3. Chemical degradation Phosphonates are very stable and breakdown of uncomplexed phosphonates requires long timescales and severe chemical conditions. At temperatures above 200 C free NTMP decomposes to various breakdown products [96,97]. These conditions are important for the fate of the chelating agents in technical systems at elevated temperatures, e.g. in cooling waters of power plants, but not for natural waters. One study performed at room temperature within the pH range of 2–10 reported that over a several month period, EDTMP hydrolyzed under formation of phosphate, phosphite and hydroxymethylphosphonate (HMP) [98]. Other phosphonate-containing breakdown products were present but were not identified. No information on the kinetics or the percentage degraded was given. In natural waters chelating agents and therefore the phosphonates always occur in the form of metal complexes. Studies on the chemical degradation of phosphonates should therefore always include the presence of metals. Degradation of the amine linkagecontaining phosphonates NTMP, EDTMP, and DTPMP was negligible in metal-ion free oxygenated solutions, but Ca, Mg, and Fe(II) brought about conversion to free phosphate at a rate of approximately 1 percent per day [99]. Although the degradation was classified as hydrolysis, the conversion rate dropped to negligible levels in the absence of O2, indicating that redox reactions play a role. HEDP, which does not contain an amine linkage, degrades approximately 20times more slowly. A loss of NTMP in different natural waters (river waters, groundwaters) and appearance of the degradation products has been observed [21]. The conversion of NTMP into iminodimethylenephosphonate (IDMP) and

B. Nowack / Water Research 37 (2003) 2533–2546

HMP was attributed to abiotic hydrolysis and the subsequent conversion to aminomethylphosphonate (AMPA) and CO2 to microbial degradation. The authors performed a follow-up study in a medium that was free of microorganisms, but contained mM levels of Ca, Mg, K, and Na and trace levels (o1 mM) of Fe(III), Cu(II), Mn(II), and Zn. Complete conversion of NTMP to IDMP, HMP and AMPA occurred within 32 h. Because multiple metal ions were present in these investigations [21,99], it was not possible to identify the catalytic agent. A systematic study on the influence of metal ions on phosphonate breakdown has been reported [33]. No breakdown of NTMP was observed in metal-free systems and in the presence of Ca, Mg, Zn, Cu(II) and Fe(III) which disagrees to previous results where degradation of NTMP was observed in the presence of Ca or Mg [21]. Very rapid degradation of aminopolyphosphonates occurred in the presence of Mn(II) and molecular oxygen [33]. The half-life for the reaction of NTMP in the presence of equimolar Mn(II) and in equilibrium with 0.21 atm O2 was 10 min at pH 6.5. The reaction occurs more slowly under more alkaline or acidic conditions. In the absence of oxygen no reaction took place, indicating that an oxidation step was involved. The presence of other cations such as Ca, Zn, and Cu(II) can considerably slow down the reaction by competing with Mn(II) for NTMP (Fig. 6). Catalytic Mn(II) is regenerated by oxygen in cyclic fashion as the reaction takes place. The hypothesized pathway is that Mn(II)-phosphonate is oxidized by molecular oxygen to the Mn(III)-phosphonate. In an intramolecular redoxreaction the Mn(III) oxidizes the phosphonic acid and is in turn reduced to Mn(II). Formate, orthophosphate, IDMP and FIDMP breakdown products have been identified. Breakdown also

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occurs in oxygen-free suspension of the Mn(III) containing mineral manganite (MnOOH) and with MnOOH in the presence of oxygen [100]. EDTMP and DTPMP are also degraded in the presence of Mn(II) and oxygen, although at a slower rate, but not the amine-free HEDP [33]. Two of the breakdown products of NTMP, IDMP and FIDMP, have been detected in WWTP [34]. This indicates that manganesecatalyzed oxidation of aminopolyphosphonate is likely to be an important degradation mechanism in natural waters. 5.4. Degradation during oxidation processes Phosphonates present in natural waters may be subject to oxidation and disinfection processes during drinking water treatment. No information on the behavior of phosphonates during chlorination is available. Ozonation of NTMP, EDTMP, and DTPMP resulted in the rapid disappearance of the parent compound in less than a minute [101]. 60–70% of the degraded phosphonate was found as phosphate; AMPA and phosphonoformic acid were also detected. The amine-free HEDP was degraded much more slowly with only 15% degradation after 30 min. The reaction pathway of EDTMP during ozonation is equivalent to that of EDTA [102]. The herbicide glyphosate was formed during ozonation of EDTMP with concentration of up to 10 nM [103]. The environmental fate, behavior and analysis of both AMPA and glyphosate has received considerable attention [10] and the formation of these compounds during ozonation of an aminopolyphosphonate may change the risk analysis of these compounds considerably.

6. Speciation

% of inital NTMP

120 100 80 60 40

no oxygen only Mn(II) Mn(II)/ 0.5 mM Ca Mn(II)/ 10 µM Zn

20 0 0

100

200

300

400

500

time (minutes) Fig. 6. Oxidation of 10 mM NTMP in the presence of 10 mM Mn(II) in the presence and absence of dissolved oxygen and competing metal ions at pH 7.0. Reprinted with permission from [33]. Copyright (2000) American Chemical Society.

The speciation of chelating agents in the environment can be calculated based on the known stability constants of the metal–ligand complexes and the measured total concentrations of metals and chelating agents. This approach has been used to predict the speciation of EDTMP in Rhine water [6]. The simulated speciation was dominated by CuEDTMP and ZnEDTMP. HEDP was predicted to be mainly complexed with Ca and NTMP with Cu and Zn [104,105]. But how accurate are such calculations? There are several points to consider: In speciation calculations it is always assumed that equilibrium has been reached in the system. This is not always the case. Some metal complexes of aminocarboxylates have very slow exchange kinetics [106]. It has been found for example that Fe(III)EDTA is not in equilibrium with other metals in river water due to slow exchange kinetics of Fe(III)EDTA [107]. Almost nothing is known about the exchange kinetics of metal-phosphonate

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complexes and therefore all equilibrium calculations have to be treated with care. Most calculations also do not consider that besides the chelating agent of interest other chelating agents and natural ligands are present in the water and compete for available metals. The interaction between the binding properties of phosphonates and fulvic acids is weak [108] but it has been shown that considering the natural ligands for Cu and Zn is critical for obtaining an accurate speciation of chelating agents [109]. In the following section a speciation model for three phosphonates is developed, based on a river water sample from Switzerland with well-known composition of metals, anthropogenic and natural ligands [110]. These ligands compete with the phosphonates for the same metals and have to be included in the speciation calculation. The concentration of the phosphonates in the calculations was set to 20 nM, comparable to EDTA at that location. The speciation was calculated for HEDP and NTMP with the constants from the IUPAC report [12] and for DTPMP with the constants from [111]. If only total metals and the phosphonates are taken into consideration, speciation is dominated by Cu for DTPMP, Ca for HEDP, and Ca, Mg, Zn and Cu for NTMP. Including EDTA and NTA does not change the speciation significantly; however, as soon as the natural ligands for Cu and Zn are considered, the calculated speciation for NTMP and DTPMP changes drastically. For NTMP the Cu and Zn complexes disappear totally due to the very strong binding of Cu to the natural ligands and CaNTMP and MgNTMP are dominant. For DTPMP the Ca and Mg complexes also become very important with more than 60% of the DTPMP complexed by these metals. CuDTPMP is only a minor species under these conditions. For HEDP the alkaline earth metals Ca and Mg are the major bound metals under all conditions. The fraction of other metal complexes is never above 0.1%. It can be concluded that phosphonates are most probably complexed to alkaline earth metals in natural waters. This calculation shows that considering the natural ligands is crucial for obtaining a reasonable result for phosphonate speciation (Table 2). Analytical methods have been developed to determine directly the speciation of aminocarboxylate chelating agents [112–114]. In principle these methods should also be applicable to phosphonates. A recent very promising method uses anion-exchange chromatography coupled to ICP-MS for the separation of metal-chelating agent complexes [35,36]. The method is also applicable to phosphonates and it has been shown that the CuEDTMP complex can be determined. The use of these methods to determine the speciation of phosphonates in natural waters is needed.

Table 2 Calculated species distribution of HEDP, NTMP, and DTPMP in river water. Conditions: 20 nM phosphonates, 29.4 nM EDTA, 8.6 nM NTA and natural ligands for Cu, Zn and Ni Ca

Mg Zn

Cu

% of total phosphonate HEDP No other ligands With EDTA, NTA, natural ligands NTMP No other ligands With EDTA, NTA, natural ligands DTPMP No other ligands With EDTA, NTA, natural ligands

88 88

12 12

0 0

0.1 0

33 55

25 42

11 0

28 0

0 41

0 21

24 35

76 2

7. Behavior during wastewater treatment The studies about the behavior of phosphonates during wastewater treatment can be divided into two groups: field studies with the addition of elevated concentrations of phosphonates to the influent of the treatment plant and investigations at ambient concentrations. The elimination of phosphonates during wastewater treatment was found to be very high, even with high concentrations of added phosphonates of about 10 mM. Elimination of 9.7 mM HEDP in a field experiment was about 60% during the sedimentation and 90–97.5% during the biological step with simultaneous FeCl3 precipitation [60]. Lower removal rates of 50–60% were found with the addition of 5–10 mM HEDP and 3–7 mM NTMP to a WWTP without iron-addition [115]. The behavior of 4.5–12 mM DTPMP was followed through the different treatment steps [39]. It was found that the DTPMP removal in the biological step was 95%. After the precipitation step with aluminum sulfate about 97% of the added DTPMP had been removed. This investigation has shown that even without simultaneous addition of iron or aluminum salts, very good removal in the biological step can be achieved. The second group of studies investigated the fate of phosphonates that are already present in the influent of the WWTP. For a 13-day field study a total amount of 117 mol of DTPMP was found in the influent of the WWTP compared to an effluent load of 17 mol, meaning that the removal efficiency was 85% [38]. Elimination of NTMP and EDTMP from another WWTP was at least 80% and 70%, respectively [38]. Because the concentration in the effluent was below the detection limit, this removal efficiency is the lower limit.

B. Nowack / Water Research 37 (2003) 2533–2546

The fate of NTMP was followed in another WWTP receiving wastewater from textile industry [39]. The load of NTMP in the influent was 324 mol during the 2-week period. No NTMP was detected in the effluent. Taking the detection limit of 0.05 mM as the upper limit of NTMP concentration the maximal effluent load for the 2-week period can be calculated to be 23 mol. The removal efficiency of the WWTP was therefore at least 93%. Also the two breakdown products of NTMP, IDMP and FIDMP, were present at much higher concentrations in the influent than in the effluent [34], with a removal of 87% FIDMP and 96% IDMP. The results from field studies and field measurements have shown that phosphonates are removed very efficiently in most WWTP and pose only little risk to the receiving waters.

8. Conclusions *

*

*

* *

*

*

The very strong adsorption of phosphonates results in low dissolved concentrations. Little or no remobilization of metals by phosphonates is expected. No biodegradation of phosphonate-chelating agents is observed in the environment. The Fe(III)-complexes are rapidly photodegraded. Rapid degradation of aminopolyphosphonates occurs in the presence of Mn(II). An analytical method for trace measurements in natural waters is urgently needed. No analytical information on speciation of phosphonates in the environment is available.

Acknowledgements The author is indebted to Jean-Claude Bollinger and V!eronique Deluchat for their fruitful comments to earlier versions of this manuscript and to Susan Tandy for editing the English. This review was prepared in part during a stay at the University of Limoges, France. The support of the Reinhold-Beitlich-Foundation, Tubingen, . Germany, is greatly acknowledged.

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