Organophosphonates: A review on environmental relevance, biodegradability and removal in wastewater treatment plants

Science of the Total Environment 615 (2018) 1176–1191

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Review

Organophosphonates: A review on environmental relevance, biodegradability and removal in wastewater treatment plants Eduard Rott a,⁎, Heidrun Steinmetz b, Jörg W. Metzger a a b

Institute for Sanitary Engineering, Water Quality and Solid Waste Management, University of Stuttgart, Bandtäle 2, 70569 Stuttgart, Germany Chair of Resource Efficient Wastewater Technology, University of Kaiserslautern, Paul-Ehrlich-Str. 14, 67663 Kaiserslautern, Germany

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Phosphonates are removed with N 80% from wastewater treatment plants (WWTPs). • Sole enhanced biological phosphorus removal may not lead to such high removal rates. • To date, phosphonates may not remobilize heavy metals from sediment significantly. • Phosphonates, except some degradation products, are harmless to most aquatic organisms. • Phosphonates contribute to eutrophication and interfere with phosphate removal in WWTPs.

a r t i c l e

i n f o

Article history: Received 10 August 2017 Received in revised form 18 September 2017 Accepted 21 September 2017 Available online xxxx Editor: D. Barcelo Keywords: Phosphonates Degradation Biodegradability Phosphate precipitation Eutrophication

a b s t r a c t The worldwide increasing consumption of the phosphonates 2-phosphonobutane-1,2,4-tricarboxylic acid [PBTC], 1-hydroxyethane 1,1-diphosphonic acid [HEDP], nitrilotris(methylene phosphonic acid) [NTMP], ethylenediamine tetra(methylene phosphonic acid) [EDTMP] and diethylenetriamine penta(methylene phosphonic acid) [DTPMP] over the past decades put phosphonates into focus of environmental scientists and agencies, as they are increasingly discussed in the context of various environmental problems. The hitherto difficult analysis of phosphonates contributed to the fact that very little is known about their concentrations and behavior in the environment. This work critically reviews the existing literature up to the year 2016 on the potential environmental relevance of phosphonates, their biotic and abiotic degradability, and their removal in wastewater treatment plants (WWTPs). Accordingly, despite their stability against biological degradation, phosphonates can be removed with relatively high efficiency (N 80%) in WWTPs operated with chemical phosphate precipitation. In the literature, however, to our knowledge, there is no information as to whether an enhanced biological phosphorus removal alone is sufficient for such high removal rates and whether the achievable phosphonate concentrations in effluents are sufficiently low to prevent eutrophication. It is currently expected that phosphonates, although being complexing agents, do not remobilize heavy metals from sediments in a significant amount since the phosphonate concentrations required for this (N 50 μg/L) are considerably higher than the concentrations determined in surface waters. Various publications also point out that phosphonates are harmless to a variety of aquatic organisms. Moreover, degradation products thereof such as N(phosphonomethyl)glycine and aminomethylphosphonic acid are regarded as being particularly critical. Despite their high stability against biological degradation, phosphonates contribute to eutrophication due to abiotic degradation (mainly photolysis). Furthermore, the literature reports on the fact that phosphonates in high

⁎ Corresponding author. E-mail address: [email protected] (E. Rott).

https://doi.org/10.1016/j.scitotenv.2017.09.223 0048-9697/© 2017 Elsevier B.V. All rights reserved.

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concentrations interfere with phosphate precipitation in WWTPs. Thus, it is recommended to remove phosphonates, in particular from industrial wastewaters, before discharging them into water bodies or WWTPs. © 2017 Elsevier B.V. All rights reserved.

Contents 1.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . 1.1. Motivation . . . . . . . . . . . . . . . . . . . . . 1.2. Physicochemical properties of phosphonates . . . . . . 1.3. Areas of application of phosphonates . . . . . . . . . 1.4. Phosphonate analysis . . . . . . . . . . . . . . . . 2. Biodegradability of phosphonates . . . . . . . . . . . . . . 2.1. C\\P bond cleavage by bacteria . . . . . . . . . . . . 2.2. Aerobic and anaerobic degradation tests . . . . . . . . 3. Phosphonate adsorption onto activated and digested sludge . . 3.1. Adsorption onto activated sludge . . . . . . . . . . . 3.2. Adsorption onto digested sludge . . . . . . . . . . . 4. Phosphonate removal in municipal wastewater treatment plants 5. Effect of phosphonates on chemical phosphate precipitation . . 6. Environmental relevance of phosphonates . . . . . . . . . . 6.1. Phosphonates in the environment . . . . . . . . . . 6.2. Environmental behavior . . . . . . . . . . . . . . . 6.3. Ecotoxicity . . . . . . . . . . . . . . . . . . . . . 6.4. Contribution to eutrophication . . . . . . . . . . . . 6.5. Remobilization of heavy metals . . . . . . . . . . . . 7. Summary . . . . . . . . . . . . . . . . . . . . . . . . . 8. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . Appendix A. Supplementary data . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . .

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1. Introduction 1.1. Motivation In nature, phosphorus is the main limiting factor for biomass growth. Thus, its removal from wastewater is a significant means to reduce the amount of nutrients entering the aqueous environment and thereby preventing algal bloom and eutrophication of water. Within the scope of the European Water Framework Directive (EU, 2000), environmental objectives have been defined which in some cases require higher phosphorus removal efficiencies in wastewater treatment plants (WWTPs) to achieve a good quality of water bodies. For this reason, for WWTPs in Europe target values of 0.1–0.5 mg/L total P (depending on the size of the WWTP) have already been established. WWTP operators can meet these more stringent requirements, among others by means of improved precipitant dosing and a more efficient secondary clarification. However, the optimization potentials are limited since the P removal efficiency depends on the composition with respect to the different kinds of phosphorous compounds present in the wastewater. With the beginning of the usage of phosphonates in the formulations of washing agents (Grohmann and Horstmann, 1989) and their application in technical processes (Reichert, 1996), the fraction of dissolved organic phosphorus (DOP) (cDOP = cdissolved P − cinorganic P; c: concentration) in the effluents of European WWTPs started to increase substantially (Neft et al., 2010; Gu et al., 2011). Phosphonates have already been identified in the influents of several German and Swiss WWTPs as part of the DOP (Fürhacker et al., 2005; Nowack, 1998, 2002b). Furthermore, due to their high tendency to adsorb onto particulate matter and solid surfaces (Nowack, 2003), phosphonates contribute to the particulate phosphorus fraction (PP) (cPP = ctotal P − cdissolved P). In

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1177 1177 1178 1178 1179 1179 1179 1180 1180 1180 1181 1182 1183 1183 1183 1185 1185 1186 1187 1188 1188 1189 1189 1189

particular, the DOP fraction is increasingly problematic for WWTP operators, as it cannot be reliably removed by P precipitation, thus making it more difficult to meet the required target values (Neft et al., 2010). The quantitatively most important phosphonates are 2phosphonobutane-1,2,4-tricarboxylic acid (PBTC), 1-hydroxyethane 1,1-diphosphonic acid (HEDP), nitrilotris(methylene phosphonic acid), ethylenediamine tetra(methylene phosphonic acid) (EDTMP) and diethylenetriamine penta(methylene phosphonic acid) (DTPMP). These compounds are versatile metal complexing agents forming complexes of high stability (Gledhill and Feijtel, 1992) and act as “thresholders” which are effective also in low, i.e., substoichiometric concentrations (Maise, 1984). Due to these properties, they are used in the paper and textile industries, for water conditioning and they are added to household and industrial cleaning products. In the early 1990s, the Europe-wide polyphosphonate consumption (excluding PBTC) was 11,820 t/a (Gledhill and Feijtel, 1992). Davenport et al. (2000), cited by Nowack (2003), estimated the phosphonate use for 1998 in Europe at 15,000 t/a and the worldwide use at 56,000 t/a. For 2012, the European Phosphonate Association indicated a European phosphonate consumption of 49,000 t/a and a worldwide consumption of 94,000 t/a (EPA, 2013). Unfortunately, EPA (2013) did not provide more detailed data on the distribution of phosphonate consumption in other continents. The enormous increase in Europe is paralleled by the global phosphonate consumption, which was 56,000 t/a in 1998 (Davenport et al., 2000) and grew by about 70% to 94,000 t/a in 2012 (EPA, 2013). With this enormous increase in consumption, phosphonates are discussed more and more from an environmental point of view and in connection with various problems such as eutrophication of water, potential remobilization of heavy metals, the formation of critical

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degradation products, and the interference with phosphate precipitation as well as the aggravating compliance with phosphorus limits on WWTPs. This review summarizes the current findings up to the year 2016 on the biodegradability and removal of phosphonates in WWTPs and gives an overview on the existing literature regarding their environmental relevance. 1.2. Physicochemical properties of phosphonates In phosphonic acids and their salts, the phosphonates, the phosphorus atom has the oxidation state +3, whereas the most frequent oxidation state for inorganic phosphorus in nature is + 5, as found, e.g., in ortho-phosphoric acid (H3PO4) or ortho-phosphates. Table 1 gives an overview of the characteristics of the most widely used phosphonates PBTC, HEDP, NTMP, EDTMP and DTPMP (structural formulae see Fig. 1). The phosphonates listed in Table 1 can be divided into two groups, the group of nitrogen-free phosphonates, which also carry characteristic carboxyl and hydroxyl groups, and into the group of aminophosphonates having three to five phosphonate groups (C–PO(OH)2). Phosphonates with more than one phosphonate group are also termed ‘polyphosphonates’. Phosphonic acids and in particular their salts have a high solubility in water (≥21 g/L), a low solubility in organic solvents (log Kow b −1.3) and a very low volatility (Henry constant b6 ∙ 10− 17 atm ∙ m3/mol) (Nowack, 2003; Gledhill and Feijtel, 1992; RSC, 2016; HERA, 2004).

and Held-Beller, 1993). Thus, the strongly pronounced pH stability of phosphonates against oxidative and thermal decomposition enable their use as bleaching stabilizer in bleaching liquors, where high temperatures prevail at high pH. • Household and industrial cleaners:

Phosphonates are used as complexing agent, scaling inhibitor and bleaching stabilizer (Nowack, 2003). • Membrane filtration:

The addition of phosphonates as antiscalants counteracts the blocking of the membrane. • Conditioning of cooling water:

Circulation cooling systems are particularly affected by scaling and related troublesome deposits of precipitates such as calcium carbonate and calcium phosphate due to the intended evaporation of the cooling water (Gartiser and Urich, 2002). Phosphonates can substoichiometrically complex “free” calcium and magnesium ions (“threshold effect”) and thus increase the degree of thickening. • Building materials industry:

1.3. Areas of application of phosphonates Due to their stability and threshold effectiveness, there is a wide range of technical applications of phosphonates. Phosphonic acids have been chemically synthesized for N70 years, but have gained importance only since the beginning of the 1980s. Since then they have increasingly replaced the highly controversially discussed aminopolycarboxylic acids, such as ethylenediaminetetraacetic acid (EDTA) and diethylenetriaminepentaacetic acid (DTPA). In the following, the areas of application of the five phosphonates PBTC, HEDP, NTMP, EDTMP and DTPMP are described: • Textile and paper production:

Native textile fibers are often contaminated with iron hydroxides and iron oxides. In addition, iron from stainless steel appliances can dissolve and deposit as hydroxide on the product. Iron, manganese and copper ions interfere with the bleaching and dyeing process by causing the catalytical decomposition of the bleaching agent with negative effect on the degree of whiteness of the treated textiles and causing catalytic damage to the substrate. Phosphonates complex metals in the alkaline pH range and thus limit the precipitation of metals in the form of metal hydroxides. For example, at pH 10, despite the high hydroxide concentration, one mol of DTPMP still binds three moles of Fe (Bachus

At high temperatures, the curing of the cement is delayed by phosphonate addition (Knepper, 2003). • Cosmetics:

Phosphonates serve, for example, as stabilizers for bleaching agents in hair bleaching (Maise, 1984). • Electroplating:

In the degreasing bath, the formation of interfering hydrophobic lime soaps, i.e., water-insoluble calcium salts of fatty acids, can occur when saline water is used (Reichert, 1996). Phosphonates complex interfering calcium cations and thus inhibit this undesirable process. • Medicine:

The use of HEDP and EDTMP has been approved for the treatment of bone diseases (EPA, 2013; de Klerk et al., 1992; Wilky and Loeb, 2013). • Oil production:

Table 1 Physicochemical properties of phosphonates (aRSC, 2016; bJaworska et al., 2002, and sources cited herein; cGledhill and Feijtel, 1992; dHERA, 2004).

CAS-No. Molecular formula Molar mass Phosphorus content (%) Phosphonate groups Amine groups Water solubility (g/L) KHenry (atm∙m3/mol) log Kow

PBTC

HEDP

NTMP

EDTMP

DTPMP

37971-36-1 C7H11O9P 270.13 11.5 1 0 N100a 3.3 ∙ 10−17 a −1.36a

2809-21-4 C2H8O7P2 206.03 30.1 2 0 N100b 5.2 ∙ 10−17 a −3.49c

6419-19-8 C3H12NO9P3 299.05 31.1 3 1 N100b 8.1 ∙ 10−18 a −3.53c

1429-50-1 C6H20N2O12P4 436.12 28.4 4 2 21c 1.2 ∙ 10−17 a −4.10c

15827-60-8 C9H28N3O15P5 573.20 27.0 5 3 N100b 7.3 ∙ 10−18 d −3.40c

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Fig. 1. Structural formulae of widely used phosphonic acids (based on ACS, 2016).

The use of scaling-inhibiting phosphonates reduces the stone formation in oil drilling (Maise, 1984). 1.4. Phosphonate analysis The knowledge of problems with regard to the determination of phosphonates in the low concentration range in samples with highly pronounced organic and inorganic contamination is of great importance, as very little information on the presence of phosphonates in the environment is available due to this fact. This section therefore summarizes briefly the development of phosphonate analysis and its challenges. The analysis of phosphonates is particularly difficult due to their low reactivity and lack of chromophores absorbing in the UV and IR range (Knepper, 2003). Various techniques for the analysis of phosphonates are presented in the literature (Nowack, 1997, 2002a; Esser et al., 2007; Fürhacker et al., 2005; Thompson et al., 1994; Klinger et al., 1997; Meek and Pietrzyk, 1988; Weiß and Hägele, 1987; Tschäbunin et al., 1989a, 1989b, 1989c; Vaeth et al., 1987; Wagner et al., 1990; Tewari and van Stroe-Biezen, 1997; Wong et al., 1987; Shamsi and Danielson, 1995; Ammann, 2002; Schmidt et al., 2014), most of which cannot be applied to environmental or wastewater samples due to a high susceptibility to interferences to matrix effects (cations, anions, organics, etc.) and high limits of quantification (LOQ N 1 mg/L). The first systematic analyses of the phosphonate concentrations in municipal wastewater samples (influent and effluent) of WWTPs were presented by Nowack (1998, 2002b). The method is based on a precomplexation of the phosphonates with FeIII followed by the photometric detection (at 260 nm) of the Fe-phosphonate complexes separated by means of ion-pair high-performance liquid chromatography (HPLC) (Nowack, 1997). This method allows the detection of aminophosphonates with limits of detection (LODs) between 0.05 and 0.5 μmol/L (15–57 μg/L for aminophosphonates). A disadvantage is that individual phosphonates are identified solely by means of the retention time (Knepper, 2003); in addition HEDP is detectable only at concentrations N100 μg/L. If several phosphonates are present simultaneously, there is also the possibility that the competition for FeIII leads to deviations of concentrations of some phosphonates compared to the results of the complexation of individual phosphonates alone. It is also questionable whether the process can reliably be applied to

concentrates with very high water hardness. Although the process envisages the use of H+-cation exchange columns to remove interfering Ca and Mg ions, these columns reach their limits when applied to very hard wastewater samples. Fürhacker et al. (2005) later presented a method that, like the method of Nowack (1997), also provided for a separation of the noncomplexed aminophosphonates by means of an HPLC anion exchange column. The detection (via the element phosphorus) was performed by means of mass spectrometry with inductively coupled plasma (HPLC-ICP-MS), so that no precomplexation of the phosphonates with FeIII was necessary. The measurement method was applicable to municipal wastewater (influent and effluent), but had too high LOQs for these types of samples of around 100 μg/L (the LODs were between 25 and 40 μg/L). Coupled systems of ion chromatography (IC) and mass spectrometry with inductively coupled plasma (ICP-MS) are known for their high performance with regard to the highly selective and excellently sensitive determination of element species (Schmidt et al., 2014). The first attempts to determine phosphonate complexes with heavy metals with this method was described by Ammann (2002), whose measurement method did not have the objective to determine the phosphonate itself, but to determine the complexed metal species (Schmidt et al., 2014). Recently, Schmidt et al. (2014) suggested a phosphonate detection method using the IC-ICP-MS principle, which combines the advantages of low LOQs for polyphosphonates in the range of 0.1 μg/L (compared to the LOQs in studies of Nowack, 1997, of at least 15 μg/L) with the applicability to environmental samples. The novelty of the method is the use of a strong cation exchanger for transferring the complexed phosphonates into “free” phosphonic acids, followed by the fiftyfold pre-concentration of the phosphonates by a weak anion exchanger prior to the analysis. Schmidt et al. (2014) applied the method only to river water. Accordingly, the method still has to be examined and modified for the applicability to wastewater. 2. Biodegradability of phosphonates 2.1. C\\P bond cleavage by bacteria There are several articles on the biodegradability of C\\P bonds in less complex and partly naturally occurring phosphonates (2-

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aminomethylphosphonic acid (2-AEP), aminomethylphosphonic acid (AMPA), glyphosate, etc.) (Zeleznick et al., 1963; Rosenberg and La Nauze, 1967; Kittredge and Roberts, 1969; La Nauze et al., 1970; Cook et al., 1978; Daughton et al., 1979; Ripke et al., 1987; Lerbs et al., 1990; Ternan et al., 1998; Kononova and Nesmeyanova, 2002; Martinez et al., 2010; Kamat and Raushel, 2013). However, there are only a few studies on the phosphonates PBTC, HEDP, NTMP, EDTMP and DTPMP, which are used as complexing agents. These industrial chemicals are not formed naturally. From an earth-historical perspective, microorganisms have been confronted with these compounds only for a very short time and thus the enzyme repertoire for efficiently degrading these substances is widely missing. Schowanek and Verstraete (1990) examined seven different environmental samples (peat soil, water from an oligotrophic lake, two samples of activated sludge, river water, aerobic compost, digested sludge) on microorganisms that are able to convert natural and xenobiotic phosphonates as the sole P source for their growth. In this respect, 19 pure bacterial cultures exemplarily selected were examined, as well. 2-AEP was best utilized by these strains. Only one strain out of these 19, Arthrobacter sp. strain GLP-1, was capable of degrading all four phosphonates HEDP, NTMP, EDTMP and DTPMP. The majority of the enriched cultures from the environmental samples could solely convert one of the phosphonate groups from the polyphosphonates. However, in a few enriched cultures, almost a complete C\\P bond cleavage in HEDP and EDTMP was observed. DTPMP could not be completely degraded by neither strain. NTMP was not considered in these studies. The following average proportions of phosphonate-utilizing microorganisms isolated from the environmental samples were determined by comparing the number of each colony forming unit in nonselective and selective nutrient medium: 2-AEP: 24.33%, glyphosate: 6.56%, HEDP: 0.18%, EDTMP: 2.49%, DTPMP: 1.31%. Thus, glyphosate as the sole P source could be utilized more easily than the polyphosphonates, which were degraded by comparatively only few microorganisms. Environmental samples, which can be expected to usually be contaminated more frequently with phosphonates, did not contain significantly more polyphosphonate-utilizing microorganisms than samples without polyphosphonates. Thus, the authors concluded that phosphonateutilizing microorganisms are ubiquitous. Raschke et al. (1994) investigated the aerobic microbial utilization of PBTC using enriched cultures of microorganisms from eight different samples (4× activated sludge, 2× river sediment, 2× river water) and the anaerobic microbial utilization of PBTC using three different environmental samples (2× river sediment, digested sludge of a WWTP), respectively. In all cultures, PBTC could not be identified as a carbon source. In the presence of alternative carbon sources (e.g., sodium acetate, ethanol) and offered as the sole P source, PBTC was degraded in all enrichment cultures. The following five strains were identified as phosphonate-utilizing strains: Acinetobacter spec. DNA group II, Acinetobacter baumannii, Alcaligenes spec., Pseudomonas spec., Methylobacter spec. These pure cultures were also unable to utilize PBTC as the sole carbon source, but converted N90% of 0.2 mmol/L of the compound within 200 h in the presence of alternative carbon sources. The authors showed that Methylobacter spec., with sufficient latency (ca. 50 h), completely utilized PBTC in an environment poor in phosphate and in the presence of pyruvate, although increasing phosphate concentrations were observed during the reaction time. Apparently, the availability of inorganic phosphorus is no longer limiting the degradation of PBTC after the enzymes necessary for C\\P bond cleavage (phosphonatase, C–P-lyase) have already been induced. Based on this finding, the authors concluded that PBTC can be degraded also in a natural environment where similar conditions prevail. The PBTC-utilizing microorganisms could not be assigned to a specific bacterial family. Slow degradation took place both in the aerobic and in the anaerobic environment, with increased conversion rates in the presence of oxygen. Schowanek and Verstraete (1990) and Raschke et al. (1994) thus agreed to the finding that PBTC- and polyphosphonate-degrading

microorganisms are ubiquitous. However, these organisms are only present in very small proportions in environmental samples. Moreover, phosphonate degradation would preferably take place in a phosphatefree environment and in the presence of suitable carbon sources. 2.2. Aerobic and anaerobic degradation tests Table 2 contains a summary of results of OECD degradation tests applied to phosphonates. The Zahn-Wellens test (OECD 302 B, ISO 9888) provides information on the inherent biodegradability of the tested chemical substance. None of the tested phosphonates was inherently biodegradable. Elimination rates measured in Zahn-Wellens tests were generally attributed to adsorption onto the inoculum. Only the results of Reichert (1996) hint to biological degradation of NTMP and EDTMP to an increased extent, since primary substance decline rates of 50–80% were observed while at the same time very low DOC removal rates were observed. This result also proves that phosphonates are not completely mineralized to CO2 and phosphate, not even after several weeks (the typical duration of Zahn-Wellens tests is 28 days). The degree of CO2 formation of b10% observed in the SCAS test (OECD 302 A, ISO 9887) by Saeger et al. (1978) substantiates this finding. Similarly, the very low extents of removal in the modified OECD screening tests (OECD 301 E, ISO 7827), the closed bottle tests (OECD 301 D, ISO 10707), the CO2 evolution test (modified Sturm test, OECD 301 B, ISO 9439) and the manometric respiration tests (OECD 301 F, ISO 9408) are a clear indication of the strong stability of the tested phosphonates against aerobic degradation. In the studies carried out under anaerobic conditions by Steber and Wierich (1986, 1987), b4% of HEDP and NTMP were converted into digester gas (CO2, CH4) within 28 days. On the other hand, however, an increased adsorption on sludge was observed (only 4–6% of the initial phosphonate concentration was found in the supernatant). Nowack (1998) also did not find any anaerobic degradability of the polyphosphonates HEDP, NTMP, EDTMP and DTPMP complexed with CaII. Furthermore, 0% of 20 mg/L Na4PBTC was degraded (related to DOC) by anaerobic sludge within 56 d (OECD SIDS PBTC, n.d.). The vast majority of the tests in which the formation of CO2 or the actual concentration of the primary substance, i.e. phosphonate, was analyzed (Reichert, 1996; Saeger et al., 1978; Steber and Wierich, 1986, 1987; Schöberl and Huber, 1988) thus agree that phosphonates are not significantly mineralized by activated sludge bacteria even within several weeks of exposure. Only Reichert (1996) found significant mineralization of aminophosphonates. In exceptional cases, the degradability can therefore depend on the individual activated sludge sample. However, Schowanek and Verstraete (1990) had shown that usually no increased degradation of phosphonates by activated sludge samples whose biocenosis should be adapted to phosphonates over several years in comparison to other environmental samples of bacteria is observed (Section 2.1). All sources, including Reichert (1996), either way agreed that the nitrogen-free phosphonates PBTC and HEDP are not or only insignificantly mineralized by activated sludge bacteria. Due to the very slow degradation and the small number of phosphonate-utilizing species of microorganisms, it is thus assumed that phosphonates cannot be mineralized in conventional biological WWTPs, neither aerobically nor anaerobically. Accordingly, phosphonates remain in adsorbed form in the sewage sludge in their original form and can only be completely oxidized (‘mineralized’) and thereby removed from the environment by combustion of the sewage sludge. 3. Phosphonate adsorption onto activated and digested sludge 3.1. Adsorption onto activated sludge Results of experiments on the adsorption behavior of phosphonates onto activated sludge are shown in Table 3. Reichert (1996) used relatively high phosphonate initial concentrations (50 and 500 mg/L) and

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Table 2 Elimination of phosphonates in OECD tests (in %). OECD test

Analytical parameter

302 B

COD DOC DOC DOC DOC DOC DOC Primary substance DOC 14 CO2/C n.i. DOC 14 CO2/C DOC BOD30/COD BOD30/COD O2 consumption O2 consumption CO2 evolution

302 A

301 E

301 D 301 F 301 B

PBTC

HEDP

NTMP

EDTMP

DTPMP

Reference

– – – – – 20 50 65 – –

50 – – – 50 – 0 30 0 3.5 –

– – – – – – –

– – – – – – –

Held (1989) Bachus (2003) Bachus (2003) Steber and Wierich (1986, 1987) Cegarra et al. (1994) BADSR (1989) Reichert (1996) Reichert (1996) Horstmann and Grohmann (1988) Saeger et al. (1978) Bachus (2003) Horstmann and Grohmann (1988) Steber and Wierich (1986, 1987) AAW (1984) BADSR (1976) Steber and Wierich (1986, 1987) Huber (1975) Metzner and Nägerl (1982) Schöberl and Huber (1988)

60–80a 20b 45c 23 – – 5 80 85 2.0 90 No biodegradability b3 20 – 0 – – b10 b38 No consumption – – b10 b10

– – – – 17 0 0 b20 – –

– – 33 – – 0 0 100 6.7 –

– – 0 – n.c. –

OECD 302 B: Zahn-Wellens test, OECD 302 A: modified semi-continuous activated sludge test (SCAS test), OECD 301 D: closed bottle test, OECD 301 E: static modified OECD screening test, OECD 301 F: manometric respiration test (sapromat), OECD 301 B: carbon dioxide evolution test (modified Sturm test), n.i.: no information, n.c.: no consumption, COD: chemical oxygen demand, DOC: dissolved organic carbon, BOD30: biochemical oxygen demand after 30 days. a Use of technical solutions without specification of the phosphonate. b Pure mineralization. c 15% mineralization and 30% adsorption with a 20-fold sludge excess.

detected comparatively high adsorption efficiencies (N 60%) at 1.5 g/L dry substance, in particular for polyphosphonates. Steber and Wierich (1986, 1987) obtained similar results with the phosphonates HEDP (0.7–21 mg/L) and NTMP (1–300 mg/L) with 2–3 g/L dry substance. Reichert (1996) as well as Metzner and Nägerl (1982) with a maximum of 50% for PBTC found a significantly poorer adsorption affinity to activated sludge compared to the polyphosphonates. Nowack (2002b) investigated the adsorption affinity of NTMP to 1 g/L of activated sludge biomass after 2 h as a function of pH and in the presence of metals. Regardless of the complexed species (CuIINTMP, FeIII-NTMP, NTMP/1 mmol/L CaII), for an initial concentration of 1.1 mg/L NTMP the best elimination efficiency (between 80 and 100%) was observed in the pH range 3 to 5. In the pH range between 6 and 9, the elimination efficiency ranged from 40 to 60%. Furthermore, Nowack (2002b) found a better adsorption of NTMP at pH 6.5 with increasing CaII concentrations. A large number of the studies discussed here and presented in Table 3 were carried out with initial phosphonate concentrations of higher than 50 mg/L, which are not very representative for WWTP influents. All in all, however, the results suggest that, despite the lack of biodegradation (see Section 2), in particular, a removal of

polyphosphonates via adsorption may occur in WWTPs. It is also striking that compared to other phosphonates PBTC is the most difficult phosphonate to eliminate. The exact mechanism of adsorption is still unclear, as most of the publications lack information on whether the activated sludge used was obtained from WWTPs operating with simultaneous precipitation, i.e., whether the activated sludge contained a significant quantity of metal hydroxides or metal oxides to which the phosphonates could have been preferentially adsorbed (Nowack and Stone, 1999; Zenobi et al., 2005). The results of Nowack (2002b) show that the adsorption depends strongly on the polarity of the surface. With CaII adsorbed, the polarity of the surface becomes more positive. This results in reduced charge repulsion to the phosphonate, so that NTMP can adsorb via ternary complexes (Stone et al., 2002). However, it is still unclear whether this effect of CaII increasing the adsorption of NTMP is also valid for other phosphonates. 3.2. Adsorption onto digested sludge The abovementioned results suggest a relatively high extent of removal of phosphonates by adsorption onto activated sludge under aerobic conditions in the wastewater treatment process. As a result, being

Table 3 Removal efficiencies of phosphonates in sludge adsorption experiments (in %). Phosphonate (mg/L)

DS (g/L)

PBTC

HEDP

NTMP

EDTMP

DTPMP

Reference

Adsorption experiments with activated sludge: 500 1.5 50 1.5 5.0–200 n.i. 0.7–21 2–3 1.0–300 3.0 1.1 1.0 1.1 1.0

40a 50b 40–50 – – – –

60a N90a – N90a – – –

60a N95a – – N90 80–100d 40–60e

45a N90a – – – – –

70a N90c – – – – –

A A B C D E E

Adsorption experiments with digested sludge: 100 1.8

10c

65c

85c

N95c

80c

A

A: Reichert (1996), B: Metzner and Nägerl (1982), C: Steber and Wierich (1986), D: Steber and Wierich (1987), E: Nowack (2002b), DS: dry substance, n.i.: no information. a After 1 day. b After 1 week. c After 1 h. d pH 3–5. e pH 6–9, in the presence of 0.5–2 mmol/L CaII: N60% removal rates.

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concentrated in the primary or secondary sludge, phosphonates are fed into the anaerobic digestion process. Consequently, Reichert (1996) investigated whether this change of milieu may lead to changes in the adsorption behavior, possibly even to desorption processes. For this purpose, activated sludge from a municipal WWTP (fresh sludge, 4.5 g/L dry substance), which had been spiked with phosphonate (100 mg/L), was aerated for one week and then mixed with digested sludge (ratio of fresh sludge to digested sludge 1:5). The filtrate was investigated at various times to determine any desorption/dissolution under anaerobic conditions at 35 °C. It turned out that polyphosphonates that were adsorbed onto activated sludge did not desorb during the anaerobic digestion process. Thus, polyphosphonates which have already been removed from the wastewater during the aerobic treatment process do not return to the WWTP via the surplus water. The monophosphonic acid PBTC, on the other hand, exhibited an accumulation tendency with ca. 30% re-dissolution. Reichert (1996) also investigated the adsorption behavior of phosphonates onto digested sludge under anaerobic conditions and found quite high differences in the behavior of the phosphonates tested (see Table 3). b20% of NTMP, EDTMP and DTPMP were detected in the filtrate at the end of the experiment of 1 h. Approximately 90% of PBTC and 35% of HEDP, however, were still found in the filtrate indicating that the monophosphonic acid could be problematic in WWTPs. Thus, so far, there are only few studies on the adsorption of phosphonates onto digested sludge. However, these few studies indicate that phosphonates adsorbed on activated sludge remain adsorbed in the anaerobic environment. Why PBTC, which apparently has the lowest adsorption tendency of all five phosphonates anyway, tends to desorb more strongly, is still unclear. PBTC is a phosphonate with only one phosphonate group, three carboxyl groups and has smaller complex formation constants in comparison to aminophosphonates (Knepper, 2003). This complex binding capacity is obviously a decisive factor for the adsorption of phosphonates. Furthermore, complexing agents based on carboxyl groups (e.g. ethylenediaminetetraacetic acid) usually have a lower adsorption affinity towards activated sludge than complexing agents based on phosphonate groups (Fischer, 1991, 1992; Bucheli-Witschel and Egli, 2001). 4. Phosphonate removal in municipal wastewater treatment plants In a field test described by Hoelger et al. (2008), DTPMP was dosed into the grinding shop of a paper production for three days and a balance of the phosphorus loads from the grinding shop over the effluent of the paper company to the end of a biologically operated WWTP was made (no concentration data). A small portion of the dosed DTPMP was already retained by adsorption onto the paper in the paper machine. Further additional adsorption processes resulted in 70 to 80% of the phosphonate still entering the WWTP. There, N90% of the incoming DTPMP was eliminated. Metzner (1990) carried out field tests in two WWTPs, the influents of which were spiked with 2 mg/L HEDP or NTMP. Neither in the WWTP with an oxidation ditch nor in the WWTP with an activated sludge process, negative effects on the BOD5, COD and DOC elimination were found. The elimination efficiency of HEDP in the oxidation ditch was between 25 and 70%, on average ca. 50%. A similar result was observed for the activated sludge process. 60% of NTMP was removed by this method, whereas the largest proportion was removed in the primary treatment. Müller et al. (1984) determined a higher removal efficiency for HEDP. At a spiked concentration of 2 mg/L in the influent of a WWTP, a removal of 50 to 70% already occurred in the primary sedimentation. In the aeration tank with simultaneous phosphorus precipitation, the remaining phosphonate was almost completely eliminated from 90 to about 97.5% (with regard to the influent concentration). The authors predominantly attributed the removal to the adsorption of phosphonate on freshly precipitated iron hydroxide flocks.

Nowack (2002b) dosed DTPMP into the influent of a Danish WWTP with a load of 7.6 kg/d. Depending on the flow rate, concentrations of 2.5 to 6.9 mg/L of DTPMP were obtained in the influent stream. The biological treatment resulted in a 95% removal of this compound. In a further step, by the addition of aluminum sulphate, the removal efficiency was increased to 97%. To date, the non-availability of a sensitive phosphonate analysis method (see Section 1.4) has led to the fact that only a few aforementioned field experiments with predominantly high phosphonate concentrations in the still measurable range of mg/L in the WWTP influent were carried out in order to determine the phosphonate removal performance of WWTPs. However, these comparatively high phosphonate concentrations are unlikely in WWTPs. In the following, the few known publications investigating phosphonate concentrations in WWTPs without spiking beforehand are described (Nowack, 1998, 2002b; Fürhacker et al., 2005). Nowack (1998) investigated the inflow and outflow concentrations of the polyphosphonates NTMP, EDTMP and DTPMP from seven Swiss WWTPs. The WWTP Herisau received discharges from the textile industry. NTMP was found in concentrations above the LOD (15 μg/L) in five WWTPs. With a week-average concentration of 0.25 mg/L (equivalent to 0.08 mg/L P), the WWTP Herisau had the highest value here. EDTMP was found in the inflow of six WWTPs. The highest average concentration found was 0.09 mg/L EDTMP (equivalent to 0.03 mg/L P). In all WWTPs, these two phosphonates could be removed to concentrations below the LOD (NTMP: 15 μg/L, EDTMP: 21 μg/L). DTPMP was detected in the inflow of three WWTPs in concentrations above the LOD (57 μg/L). The WWTP of Herisau had an average concentration of 0.97 mg/L DTPMP, which corresponds to a phosphorus content of 0.27 mg/L. The example of the WWTP in Herisau shows that phosphonate concentrations of N 1.25 mg/L (corresponding to 0.4 mg/L P) can be present in WWTP inflows. WWTPs with predominantly domestic wastewater had only very low phosphonate concentrations. The question whether the use of precipitants contributed to an increased elimination efficiency could not be answered since the wastewater of the only WWTP without phosphate precipitation had too low phosphonate concentrations. When the WWTP of Herisau was examined in more detail, a typical weekly course of the DTPMP concentration with low concentrations at the weekend was determined. NTMP was removed by at least 80% and EDTMP by at least 70% in all WWTPs. All samples were membrane-filtered (0.45 μm pore size) prior to the analysis. Thus, any phosphonates adsorbed onto solids in the influent as well as in the effluent were not taken into account. Furthermore, Nowack (2002b) analyzed the NTMP loads in the inflow and outflow of a German WWTP operated with simultaneous chemical precipitation (Weil) with multiple indirect dischargers of the textile industry. Here, the NTMP concentration in the inflow varied between 60 μg/L and 0.3 mg/L. Occasionally, NTMP loads of 18 kg/d in the inflow were achieved. The two-week survey ultimately resulted in an NTMP removal efficiency of at least 93%. Fürhacker et al. (2005) analyzed the concentrations of aminophosphonates in the influent and effluent of five WWTPs. EDTMP was detected above the LOD of 29 μg/L (LOQ: 125 μg/L) only in the feed of the two WWTPs with the highest textile industry share (167 and 49 μg/L EDTMP). In these WWTPs, EDTMP was removed to concentrations below the LOD. NTMP (LOD: 35 μg/L) and DTPMP (LOD: 38 μg/L) were not detected. In all publications but one (25–70% removal efficiency; Metzner, 1990) usually a good phosphonate removal efficiency of N 80% in municipal WWTPs is described with phosphonate concentrations in the effluent below the LODs (see also Table 4). However, relatively high LODs, like 0.057 mg/L in the case of DTPMP (Nowack, 1998), are in a range that does not rule out a potential impact of phosphonates on the eutrophication of surface waters. In particular, the concentration of all phosphonates in total can be much higher than the LOD of a single phosphonate. Furthermore, extensive studies on individual compounds such

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as the nitrogen-free phosphonates PBTC and HEDP are still missing. For example, PBTC shows the lowest tendency towards adsorption onto activated sludge in laboratory tests (Reichert, 1996). Furthermore, the very low removal efficiencies described by Metzner (1990) also suggest that some mechanical-biological WWTPs without chemical phosphorus precipitation can only achieve removal efficiencies of 25–70%. 5. Effect of phosphonates on chemical phosphate precipitation If phosphate precipitation is carried out with aluminum or iron salts on WWTPs, phosphonates, which are capable of complexing metals, play an important role on this precipitation process. Accordingly, they can contribute to increased β values (molar ratio of precipitant concentration to initial total P concentration), i.e., precipitant concentrations, as reported on in the following publications. Horstmann and Grohmann (1984, 1986) fed a test apparatus equipped with a gravel filter for flocculation filtration with formazinespiked tap water (pH 6.8–7.4). This tap water was spiked with different mixing ratios of phosphonate-P and ortho-phosphate-P. The total P concentration was constant regardless of the mixing ratio and lay between 0.6 and 2.2 mg/L depending on the individual tested phosphonate. Even with a share of only 33% HEDP of total P, the β value required for an 80% total P removal increased from 1 (in the case of 0% HEDP-P) to 5. For EDTMP, already a share of 10% of total P led to an increase of the β value from 1 to 5. Reichert (1996) carried out experiments with municipal raw wastewater (simulation of a pre-precipitation) and clarified wastewater (simulation of a post-precipitation). In the tests with raw wastewater, disturbances of the phosphate removal were already observed at concentrations of around 1 mg/L PBTC and NTMP. In the case of clarified wastewater, slightly higher concentrations of NTMP (between 1 and 10 mg/L) were required for this effect. When FeIII is used for phosphate precipitation in WWTPs, iron hydroxide is precipitating in parallel to FePO4. Phosphonates and phosphate compete for free adsorption sites of the formed flocks. Nowack and Stone (2006) modeled the simultaneous presence of NTMP and phosphate under representative conditions in WWTPs (pH 7.0, FeIII concentration of 5 mg/L) in order to estimate the phosphonate-induced reduction in phosphate removal efficiency. For example, a 10% reduction can occur in the presence of 3 mg/L phosphate and an NTMP concentration of 0.15 mg/L (which according to the authors is usually present in WWTPs). At a concentration of 1 mg/L NTMP (which is a rare, but possible phosphonate concentration; see Section 4), the reduction ranged from 10% in the presence of 9.5 mg/L phosphate to 50% in the presence of 0.95 mg/L phosphate. From this, it can be concluded that the presence of phosphonates makes it difficult to attain the desired low phosphate effluent concentrations. Wenger-Oehn et al. (2005) reported on problems with the treatment of wastewater from the beverage industry in a biological WWTP. The reduction of the total P concentration of ca. 15 mg/L (70% of it constituted phosphonates) was only possible with a high excess of

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precipitant (β value = 2.5). As a result, there was a shortage of phosphorus as essential nutrient in the biological treatment step resulting in sudden growth of filamentous bacteria and disturbance of plant operation. Furthermore, König et al. (2002) reported on a WWTP in Austria that had problems with the compliance with the phosphorus limit value of 0.5 mg/L since it had high feed loads of phosphonates, presumably via indirect discharges. An overall assessment of the influence of phosphonates on phosphate precipitation has not yet been possible. Especially for cases where industrial wastewater with elevated phosphonate loads is indirectly discharged, there are still insufficient data on phosphonate concentrations in WWTP influents. In the publications mentioned (Horstmann and Grohmann, 1984, 1986; Nowack and Stone, 2006), there are indications that phosphonate concentrations in the sub-ppm range can lead to increased β values. In order to ensure stable P removal, the consideration of individual cases of indirect discharges is therefore important. Although the removal of phosphonates can be predominantly successful, for example by adsorption onto activated sludge in WWTPs (see Section 4), this fact should not distract from a possible cost-increasing disadvantage that increased phosphonate concentrations in the influent entail. Furthermore, it is also irrelevant whether the phosphonates are adsorbed onto the activated sludge or onto metal flocks. By ternary ligand-like complex formation, i.e., the simultaneous bonding of two metal atoms by a single phosphonate as described by Stone et al. (2002), the phosphonate can also bind dissolved metals in the adsorbed state. 6. Environmental relevance of phosphonates 6.1. Phosphonates in the environment According to EPA (2013), the amount of phosphonate used in 2012 in Europa was 49,000 t, which is 52% of the global consumption of 94,000 t/a. Accordingly, Europe is the continent with the highest consumption of phosphonates. A breakdown of the consumption quantities of the individual areas of application in Europe is shown in Fig. 2 (EPA, 2013). Destatis (2013, 2015) provides information on domestic discharges and delineates the wastewater amounts of direct and indirect discharges in a sector-specific manner so that the phosphonate loads can be divided roughly into the abovementioned three wastewater streams (Fig. 2) (the data from Germany are considered to be largely representative of total Europe). It can be seen that approximately 85% of phosphonates in WWTPs can be attributed to domestic use. However, the appearingly small share from the indirect dischargers should not be neglected. Thus, in certain cases, very high proportions of wastewater from indirect dischargers, e.g., the textile industry, may lead to greatly increased phosphonate concentrations in influents of WWTPs (Neft et al., 2010; Nowack, 1998, 2002b). A significant proportion of the phosphonate load in water is obtained by means of membrane concentrates polluted with antiscalants as well as cooling waters for which pretreatment is not standard (Gartiser and Urich, 2002).

Table 4 Removal efficiencies of phosphonates in municipal WWTPs (in %). Phosphonate in the influent (mg/L) a

n.i. 2.0a 2.0a 2.5–6.9a 0.06–0.3b Max. 0.25b Max. 0.07b Max. 0.97b 0.167b

PBTC

HEDP

NTMP

EDTMP

DTPMP

Reference

– – – – – – – – –

– 25–70 N90 – – – – – –

– 60 – – N93 N80 – – –

– – – – – – N70 – N80

N90 – – 95–97 – – – 85 –

A B C D D E E E F

A: Hoelger et al. (2008) B: Metzner (1990) C: Müller et al. (1984) D: Nowack (2002b) E: Nowack (1998) F: Fürhacker et al. (2005) n.i.: no information. a Spiking of phosphonate. b Without spiking of phosphonate.

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Fig. 2. Phosphonate loads of direct dischargers and municipal WWTPs in Europe based on values from the years 2012 and 2013 (EPA, 2013; Wiechmann et al., 2012; Nowack, 1998, 2002b; Destatis, 2013, 2015; Eurostat, 2017a; Schmidt et al., 2014).

According to Fig. 2, the approximate phosphonate load in the inflow of European WWTPs is 24,600 t/a (7400 t/a phosphonate-P; calculations based on EPA, 2013; Destatis, 2013, 2015; neglecting the connection rate; see Supplementary data). Considering a phosphonate removal efficiency of 80–95% in municipal WWTPs mainly by adsorption onto activated sludge (Section 2.2 and Section 4), approximately 19,700–23,400 t/a phosphonate is adsorbed onto sewage sludge. According to Eurostat (2017a), ca. 15 million tonnes of sewage sludge dry substance are generated annually on European municipal WWTPs. Accordingly, the average phosphonate loading of sewage sludge is 1.2–1.5 g of phosphonate/kg of dry substance. In Europe, approximately 37% of the sewage sludge is used in agriculture as fertilizer and for soil improvement (Wiechmann et al., 2012). Thus, approximately 7300–8700 t/a phosphonate is discharged via sewage sludge in agriculture. Phosphonates have a very low mobility in soils (Steber and Wierich, 1986, 1987). This explains why the risk of contaminating the groundwater with phosphonates via the application of phosphonate contaminated sewage sludge appears to be rather low (Gledhill and Feijtel, 1992). Notwithstanding this finding, Knepper (2003) critisized the lack of a comprehensive risk assessment regarding the input of phosphonates via sewage sludge applied in agriculture. The phosphonate load discharged into receiving water by municipal WWTPs in Europe can be estimated at 1200–4900 t/a at an assumed

removal efficiency of 80–95% (Section 4). These numbers can be even higher since the calculations do not take into account the low connection rates of WWTPs that still are fact in some European countries such as Romania and Albania (Eurostat, 2017b). A selective phosphonate removal is not yet state of the art in the industries directly discharging their effluents. Thus, a very rough estimate of 7800–13,700 t/a can be given for the phosphonate load discharged by direct dischargers into receiving water. The largest share of this load is probably caused by membrane concentrates and cooling water (about 65–90%). In total, 9000–18,600 t/a phosphonate is thus discharged into receiving waters in Europe. Only recently, Schmidt et al. (2014) developed a method that allows the quantification of phosphonates in environmental samples with LOQs in the range of 0.018–0.1 μg/L. Before, there was no sensitive analytical method available (Section 1.4). Thus, due to the lack of measured phosphonate concentrations in surface water, the necessity for calculations based on models arose. In the early 1990s, Gledhill and Feijtel (1992) described a model calculation for Western European countries considering an annual polyphosphonate consumption of 6300 t (estimated consumption in the whole of Europe was 11,820 t/a). The authors estimated the concentrations of polyphosphonates in municipal WWTPs at 170 to 290 μg/L (details of calculation are missing) and assumed effluent concentrations between 90 and 235 μg/L. Considering

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dilution effects in the receiving water, accumulation in the sediment (100:1) and partial degradation by photolysis and microorganisms, they calculated dissolved concentrations of polyphosphonates in Western European waters in the order of 0.25 μg/L. Using model calculations, Jaworska et al. (2002) presented the following predicted concentrations of dissolved phosphonates in waters in the Netherlands: 0.32–4.90 μg/L HEDP, 0.07–1.15 μg/L NTMP, 0.04–0.64 μg/L DTPMP. In addition, the enrichment in sediments was calculated to be 0.08–1.18 mg HEDP/kg DS (dry substance), 0.02–0.30 mg NTMP/kg DS and 0.003–0.10 mg DTPMP/kg DS. Fürhacker et al. (2005) measured the concentration of phosphonates by means of HPLC-ICP-MS in a number of Austrian receiving waters. However, all phosphonates were below the LOQs of approximately 100 μg/L, so that based on the abovementioned estimates, a phosphonate content above this order of magnitude is unlikely in rivers. The only previously known analyses of phosphonates in surface waters with sufficiently low LOQs were described by Schmidt et al. (2014). They determined HEDP concentrations between 0.3 and 1.6 μg/L and DTPMP concentrations between 0.1 and 1.3 μg/L in six German rivers using IC-ICP-MS with metal complex cleavage and phosphonate preconcentration using ion exchangers as analytical method. NTMP, EDTMP and HDTMP (hexamethylenediamine tetra(methylene phosphonic acid)) were not quantified in any river above their LOQs between 0.095 and 0.117 μg/L. The concentration ranges of HEDP and DTPMP measured were in good agreement with the model calculations of Jaworska et al. (2002). Thus, the hypothesis of Gledhill and Feijtel (1992), that the total concentration of phosphonates in Western European waters would be usually of the order of 0.25 μg/L, was not confirmed. In all rivers examined, total phosphonate concentrations of at least 0.5 μg/L were measured. In part, the measured total concentration of phosphonates even exceeded the value given by Gledhill and Feijtel (1992) by one order of magnitude (highest total concentration in the river Pfinz in Karlsruhe, Germany: 2.9 μg/L). The concentrations reported on by Gledhill and Feijtel (1992) were calculated for the 1990s and are therefore no longer relevant because the amount of phosphonates used since then has increased considerably.

pathways of phosphonates, which are essential for understanding the behavior of phosphonates in natural waters. Despite their high solubility and due to their strong polarity, phosphonates have a very high tendency to adsorb onto sediments and suspended matter. Thus, they are considered to exist in the adsorbed state in waters (Steber and Wierich, 1986, 1987; Fischer, 1992). Phosphonates can be oxidized by oxygen when present as manganese complex (Nowack and Stone, 2000). Another possible form of degradation reaction of phosphonates is hydrolysis. Schowanek and Verstraete (1990, 1991) found light-independent hydrolytic degradation of aminophosphonates in the range of 6 to 15‰ per day measured as the conversion of organic phosphorus into inorganic phosphorus. At a rate of 0.5‰ per day, the nitrogen-free HEDP hydrolyzed only insignificantly. Different from the metal-free state, phosphonates can be easily cleaved by UV radiation in the complexed state (Matthijs et al., 1989; Grohmann and Horstmann, 1989). Thus, the degradation of phosphonates by hydrolysis is less pronounced than the metalcatalyzed degradation by photolysis (Jaworska et al., 2002). This form of phosphonate degradation takes place in the surface-near layer of water bodies, where UV radiation is not completely absorbed (Grohmann and Horstmann, 1989; Fischer, 1993). In water, only few species of microorganisms are capable of metabolizing polyphosphonates (Schowanek and Verstraete, 1990) and PBTC (Raschke et al., 1994) (Section 2.1). In particular, HEDP, with its high persistence against microbial degradation and extraordinary resistance to hydrolysis, is considered to have a lasting stability in aquatic ecosystems (Fischer, 1993). Microbial phosphonate degradation is possible both in sediment and in solution (Steber and Wierich, 1987). Furthermore, bacterial metabolism of phosphonates occurs in parallel to phosphate release by abiotic degradation mechanisms, in particular, metal-catalyzed photolysis (Gledhill and Feijtel, 1992). Accordingly, microorganisms devoid of enzymes capable of cleaving C\\P bonds can utilize abiotically formed phosphate and, thus, indirectly assimilate phosphonates and their degradation products.

6.2. Environmental behavior

In various fish studies, phosphonates were found to be active in concentrations of N100 mg/L (predominantly measured as no observed effect concentration (NOEC) or lethal concentration for 50% of the test population (LC50) after 96 h) (Knepper and Weil, 2001; Kästner and Gode, 1983; Gledhill and Feijtel, 1992; Huber, 1975; Maise, 1984; Metzner and Nägerl, 1982; Schöberl and Huber, 1988; Jaworska et al.,

Phosphonates predominantly exist in water in the form of complexes with the alkaline earth metal ion Ca2+ (Nowack, 2003) and the Fe2 + ion as predominant central ions (Jaworska et al., 2002). Fig. 3 gives an overview of the characteristic degradation and adsorption

6.3. Ecotoxicity

Fig. 3. Presence, degradation and adsorption pathways of phosphonates in natural waters (Me: metal) (with permission from Rott et al., 2017a).

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2002). Such high concentrations do not occur in natural waters (Section 6.1). However, the lethal concentration (LC) does not have any direct correlation to potential chronic effects that may already occur at much lower concentrations. Steber and Wierich (1986, 1987) observed bioconcentration factors (concentration of the test substance in the fish in relation to the concentration of the test substance in the test water) between 15 and 25 in zebrafish for HEDP and NTMP. These values are unexpectedly high for hydrophilic compounds with n-octanol/water distribution coefficients log KOW b − 1.3 (Table 1), but much lower than the bioconcentration factors of lipophilic compounds such as 1,2,4-trichlorobenzene or γhexachlorocyclohexane (lindane), which can be several orders of magnitude higher (Geyer et al., 1985, 1997). At the end of several weeks of exposure, immediately a rapid elimination of the phosphonate in the fish body occurred. These results showed that the accumulation of the phosphonate is not a result of lipophilicity, but caused by its general high adsorption affinity or other not known uptake pathways or elimination kinetics. Phosphonates were also tested in common Daphnia assays with Daphnia magna (the large water flea), where they, however, did not show any acute toxicity with EC0 values N 100 mg/L (the endpoint of these tests was the unability of the individuals to swim (immobilization) after 24 h or 48 h of exposition) (Knepper and Weil, 2001; Schöberl and Huber, 1988; Kästner and Gode, 1983; Gledhill and Feijtel, 1992; Grohmann and Horstmann, 1989). Of the invertebrates tested, oysters were most sensitive with mean effect concentrations (EC50) after 96 h of b 100 mg/L (e.g., EDTMP: EC50 = 67 mg/L) (Gledhill and Feijtel, 1992). The shell formation of oysters requires the precipitation of calcium carbonate, which is prevented by complex formation (Gledhill and Feijtel, 1992). It has to be noted that the measured mean effect concentrations (≥ 67 mg/L phosphonate) were very high compared to the naturally occuring phosphonate concentrations in the aqueous environment (Section 6.1), so that it is questionable if this adverse effect of phosphonates on oysters also occurs at these lower natural concentrations. Gledhill and Feijtel (1992) reported on studies on the algal toxicity (Selenastrum capricornutum, latterly: Raphidocelis subcapitata) of polyphosphonates. Mixtures of 10,000 algae cells/mL were exposed to various concentrations of the polyphosphonates at a light intensity of 4300 lx in algae growth medium. The chlorophyll content was analyzed via the determination of chlorophyll fluorescence. After an exposure time of 4 days, average effect concentrations between 0.4 and 20 mg/L were obtained. After 14 days, equal or significantly higher effect concentrations (i.e., lower algae toxicity) were measured for all phosphonates tested. The authors hypothesized that a photolytic degradation of the phosphonates might have occurred over the longer test period, causing that on the one hand important elements for algae growth, e.g., Ca and Mg, would be available again due to the absence of complex formation. On the other hand, additional biologically available phosphate might have been released from degraded phosphonates stimulating the algae growth. Up to phosphonate concentrations of 2.5 g/L, no clear inhibitory effect on Photobacterium phosphoreum, a light-emitting, halophilic, nonspore-forming, Gram-negative, and facultative anaerobic rod-shaped bacterium was observed (Grohmann and Horstmann, 1989). Huber (1975) investigated potential inhibitory effects of HEDP and NTMP on the aerobic biodegradation of peptone, glucose and municipal wastewater by activated sludge. NTMP had no significant effect up to the highest concentration tested (100 mg/L). HEDP had no inhibitory effect up to a concentration of 50 mg/L. In some studies, also the influence of phosphonates on the toxicity of heavy metals was investigated. Gledhill and Feijtel (1992) compared the acute toxicity (LC50) of CdII, CuII, PbII and ZnII (LC50: 0.81–73 mg/L) with the toxicity of their 1:1-phosphonate-complexes with the blue sunfish as test organism. They determined LC50 values above 100 mg/L for each of the tested metal-polyphosphonate complex. Hanstveit and

Oldersma (1996), cited by Jaworska et al. (2002), also observed no growth inhibition by complexation of heavy metals for the algae Raphidocelis subcapitata. Koch (1995) analyzed the inhibitory effect of PbII, CdII and CuII as well as their 1:1-complexes with HEDP and NTMP on the heterotrophic biocenosis of sludge of a WWTP (addition of Nallylthiourea to inhibit the nitrificants) by means of a respirometer. NTMP had no influence on the inhibitory effect of CuII. HEDP was able to reduce the effect, but not to zero. The inhibitory effect of PbII and CdII was not significantly affected by complex formation with HEDP and NTMP. In a comprehensive risk assessment study of HEDP, NTMP and DTPMP on humans and the environment (HERA, 2004), phosphonates were characterized to have low to moderate acute oral toxicity and low acute dermal toxicity. No genotoxic, mutagenic, carcinogenic nor a teratogenic effect could be detected for any of the phosphonates. Phosphonates are therefore classified as being toxicologically harmless for a wide range of aquatic species (Gledhill and Feijtel, 1992; Maise, 1984; Metzner and Nägerl, 1982; Kästner and Gode, 1983; HERA, 2004). From the mass of the degradation products of aminophosphonates, in particular, N-(phosphonomethyl)glycine (glyphosate) and aminomethylphosphonic acid (AMPA) are prominent (Klinger et al., 2000; Grandcoin et al., 2017). Glyphosate is a broad-spectrum herbicide, whereas AMPA is the main degradation product of glyphosate. AMPA is considered to have genotoxic effects (Mañas et al., 2009). Both compounds are controversially discussed in the scientific community with regard to their potential environmental relevance (de Roos et al., 2003; Eriksson et al., 2008; Hardell et al., 2011). 6.4. Contribution to eutrophication Eutrophication is the enrichment of water bodies with mineral nutrients (P and N) resulting in an excessive production of autotrophic organisms (phytoplankton) predominantly in the form of green algae and cyanobacteria (blue algae). The excessive growth of these organisms results in less light reaching the depths of the water body. The absence of light leads to the dying of the macrophytes growing on the sediment. Furthermore, phytoplankton sinks down to the sediment and is decomposed there by destruents, causing further nutrient release and phosphate mobilization from the sediment. The increased respiratory rate of the bacteria on the bottom results in oxygen deficiency (hypoxia), the increased activity of anaerobic microorganisms and the associated release of reduced toxic compounds such as hydrogen sulfide and ammonia. This self-reinforcement of eutrophication ultimately accelerates the dying of aquatic organisms. Phosphorus is the minimum factor for plant growth and eutrophication. Autotrophic organisms are only able to assimilate phosphorus in the form of ortho-phosphate (Correll, 1998). Only in very few publications, phosphonates are linked with the topic of eutrophication (Boels et al., 2010; Reichert, 1996; Horstmann and Grohmann, 1986; Studnik et al., 2015). The hypothesis that phosphonates do not contribute to eutrophication is generally attributed to the high stability of phosphonates against biodegradation by microorganisms (Metzner and Nägerl, 1982; Huber, 1975). However, studies on the biodegradability of phosphonates have so far been carried out predominantly with heterotrophic bacteria (Schowanek and Verstraete, 1990; Raschke et al., 1994) rather than with autotrophic organisms which actually are responsible for eutrophication. Furthermore, phosphonates are subject to natural abiotic degradation processes that contribute to the release of biologically available phosphate from phosphonates (Section 6.2). Gledhill and Feijtel (1992) exposed the algal species Selenastrum, Chlorella and Anabaena (cyanobacterium) (10,000 cells/mL) to different concentrations of the phosphonates NTMP, HEDP, EDTMP and HDTMP (0.1, 1.0, 10, 100 mg/L) in algae growth medium at a light intensity of 4300 lx. The chlorophyll determination by fluorescence measurement after eight days showed that NTMP stimulated the growth of

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Selenastrum and Chlorella at concentrations between 0.1 and 10 mg/L and stimulated the growth of Anabaena at concentrations between 0.1 and 1.0 mg/L. Higher NTMP concentrations caused growth inhibition of all three species. The growth of Chlorella was also stimulated by HEDP, EDTMP and HDTMP in the concentration range between 0.1 and 10 mg/L. HDTMP in all tested concentrations up to 100 mg/L caused an increased growth of the cyanobacterium Anabaena. During the first days of these tests, no growth stimulation occurred, rather the growth of the three species were inhibited. The authors ascribed the initial inhibition to the shortage of essential nutrients caused by phosphonate complexation. In the course of the experiments, these phosphonates were photolytically cleaved, which resulted in both the release of the beforehand complexed nutrients and of newly formed biologically available phosphate. Forlani et al. (2008) have shown that the cyanobacterium Spirulina platensis can catalytically cleave the C\\P bond in glyphosate. Consequently, Forlani et al. (2011) exposed Spirulina platensis to HDTMP as the sole phosphorus source and also observed a partial degradation of this phosphonate. Grohmann and Horstmann (1989), on the other hand, ruled out the bioavailability of the phosphonates PBTC, HEDP and NTMP for Scenedesmus (green algae) and Nostoc (cyanobacteria). The only previously known field trials investigating the bioavailability of phosphonates in water bodies were carried out by Grohmann and Horstmann (1989). Six square concrete tanks of 10 m3 were filled in parallel with groundwater from a large storage tank mixed with 2% effluent from a WWTP as source of nutrients (20 L/h flow rate, 21 days theoretical residence time, 49 days test duration). Three different concentrations of ortho-phosphate-P (approx. 20, 55 and 190 μg/L P) and HEDP-P (likewise) were adjusted in the tanks. Nitrate-N was dosed to each tank at the same concentration of 10 mg/L. Compressed air introduced via perforated PVC pipes on the bottoms of the tanks was used for continuous mixing, which, however, was only achievable to an unsatisfactory extent. In the course of the experiment, a strong dominance of filamentous green algae of the genus Cladophora spec. developed. At the end of the experiment, its dry mass was approximately proportional to the P concentration adjusted with ortho-phosphate in the tanks (50, 205 and 271 g of dry mass). Accordingly, the readily available phosphate was quasi completely incorporated into the biomass. In the basins with HEDP feed, to some extent also the genus Cladophora spec. developed (0, 5 and 67 g of dry mass), which was attributed to a partial uptake of the phosphorus from the phosphonate into biomass. The authors attributed this to the photolysis of the phosphonate and the successive release of biologically available ortho-phosphate. The results of such field trials are influenced by many factors. Thus, in one basin, the formation of algae biomass exceeded the theoretically possible amount (55 μg/L P: 205 g of dry mass determined, 90 g of dry mass theoretically possible). An influence of surrounding flora and fauna or an erroneous calculation of the theoretically possible algae biomass cannot be ruled out. The authors themselves found no explanation for this fact. The use of concrete tanks was also insufficient to simulate the presence of sediments or suspended matter in water bodies. Phosphonates have a high adsorption affinity for mineral surfaces such as soils and sediments (Steber and Wierich, 1986, 1987). Zaranyika and Nyandoro (1993) found for glyphosate that it is degraded more slowly in the adsorbed state than in dissolved form. According to Nowack (2003), a similar behavior can also be assumed for polyphosphonates. The adsorption on sediment and suspended matter is considered to lead to concentrations of dissolved phosphonates in surface water generally in the lower μg/L-range (Section 6.1). The total phosphorus content of phosphonates is 10 to 30% (Table 1). However, not for all water bodies a standard value can be assumed for the phosphonate concentration. The discharge of untreated cooling water or membrane concentrate contaminated with phosphonates as well as insufficiently purified municipal wastewater can contribute to phosphonate concentrations in water that deviate strongly from the predicted average

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concentration (Gledhill and Feijtel, 1992). This is relevant since even a total P concentration of as low as 20 μg/L can cause eutrophication (Correll, 1998). Despite the many factors influencing their field trials, Grohmann and Horstmann (1989) have clearly shown that, due to its photolytic decomposition, the phosphonate HEDP should be considered as a nutrient resource in a slowly flowing surface water. The growth of Spirulina platensis based on phosphonic acids (Forlani et al., 2011) as well as the growth stimulation of algae and the cyanobacterium Anabaena by HDTMP at a wide range of concentrations (Gledhill and Feijtel, 1992) speak for the transformation of phosphonates in the environment by some autotrophic organisms. Consequently, the contribution of phosphonates to eutrophication, although the process may be only slow, should not be underestimated. 6.5. Remobilization of heavy metals Complexing agents are capable of dissolving metals bound in the sediment. Consequently, remobilized metals can potentially reach toxic concentrations for aquatic organisms in water, may enter the drinking water via bank filtrate or contribute to algal bloom (HERA, 2004). Table 5 summarizes the minimum concentrations of various phosphonates at which a metal-remobilizing effect could be detected in batch extraction tests. The numbers in brackets correspond to the highest phosphonate concentration tested in the test for the case when no metal remobilization was observed. As a result of differences in the experimental conditions (pH, temperature, test duration) as well as in the kinds of sediments used, observed effect concentrations differed significantly from one another. In principle, however, it can be concluded that phosphonates have tendencies to preferably remobilize the transition metals iron, chromium and zinc. However, their remobilization only occurs at quite high phosphonate concentrations N50 μg/L which are higher than those usually found in natural waters (Section 6.1). Remobilization of the highly toxic heavy metals cadmium, lead and mercury by phosphonates occurred only at high phosphonate concentrations of N 1 mg/L and is therefore not likely to occur in water bodies (Gledhill and Feijtel, 1992; Bordas and Bourg, 1998; Nowack, 2003). The comparatively weak remobilizing effect of phosphonates on metals was explained by their high adsorption affinity for sediment (Gledhill and Feijtel, 1992). Accordingly, the phosphonates become effective only when the adsorption capacity of the sediment is exceeded. The comparison of the phosphonate concentrations necessary for remobilization of heavy metals found in batch extraction tests (N 50 μg/L) indicated in Table 5 with the likely concentrations in surface waters, which are lower than 50 μg/L (see Section 6.1), therefore allows the assessment that the risk of remobilization of heavy metals from sediment by phosphonates can be considered to be very low. Table 5 Phosphonate concentrations with remobilizing effect of metals from sediment (in mg/L). Data in brackets: Maximum concentration tested in the complete absence of metal remobilization. Substance

Fe

Cr

Zn

Cd

Ni

Pb

Cu

Hg

Tl

Ref.

PBTC HEDP

(1) 0.05 (1) 2.4 – – 0.05 0.05 – 0.05 –

(1) 0.05 (1) (2.4) – – 0.05 0.05 – 0.05 –

(1) 0.05 (1) (2.4) – – 0.05 (0.05) 0.3–1 (0.05) 0.3–1

(1) (5) (1) (2.4) 20 (100) (5) (5) 0.3–1 (5) 0.3–1

– (5) – (2.4) – – (5) (5) 0.3–1 (5) 0.3–1

(1) (5) (1) (2.4) 20 – (5) (5) – (5) –

(1) (5) (1) (2.4) 2 – (5) (5) – (5) –

– (5) – – – – (5) (5) – (5) –

– – – – – (100) – – – – –

1 2 1 3 4 5 2 2 6 2 6

NTMP EDTMP DTPMP

Ref.: reference, 1 Knepper and Weil (2001), 2 internal report from Monsanto cited by Gledhill and Feijtel (1992), 3 Müller et al. (1984), 4 Bordas and Bourg (1998), 5 Günther et al. (1987), 6 internal report from Procter and Gamble cited by Gledhill and Feijtel (1992).

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Müller et al. (1984) carried out laboratory experiments to examine the remobilization of metals from suspended activated sludge by HEDP. The metals zinc, chromium, nickel, copper, lead and cadmium were not remobilized by 2.4 mg/L HEDP. On the other hand, an increase of the solute concentration of iron was observed. Müller et al. (1984) also made a similar observation on a WWTP. The addition of 2 mg/L of HEDP to the feed of the WWTP caused significantly higher concentrations of iron in the discharge of the primary clarifier compared to the regular operation situation. However, due to the adsorptive removal of the phosphonate in the subsequent activated sludge stage, the complexed iron was removed with the adsorbed phosphonate, so that no residual concentration of iron in the wastewater was detectable. Zinc, chromium, nickel, copper, lead and cadmium did not show any unusual behavior. These findings indicate that there is no metal remobilization to be expected from activated sludge due to phosphonates, however, still more investigations are needed in order to make a sure statement. 7. Summary Articles on phosphonates with respect to their environmental relevance and degradability were published mainly in the periods 1985–1995 and 1998–2005. In the last 10 years, this subject has been rarely addressed. However, with the significant increase in phosphonate production and consumption over the last few years also the interest in their environmental behavior increased causing an increasing need for research. In the following, the findings on the state of knowledge regarding the environmental relevance of phosphonates and their removal from wastewater are summarized. Biodegradability and removal in municipal WWTPs: • Only a very small proportion of classes of microorganisms in the environment (including those constituting activated sludge) is capable of degrading PBTC and polyphosphonates (Raschke et al., 1994; Schowanek and Verstraete, 1990). • The microbial degradation of phosphonates takes place preferably in a phosphate-free environment in the presence of other carbon sources (Raschke et al., 1994; Schowanek and Verstraete, 1990). • In aerobic OECD tests, phosphonates were usually only partially mineralized, even after several weeks. At the same time, relatively high DOC removal rates, which were attributed to adsorption onto activated sludge, were also occasionally observed (Table 2). • Phosphonates were not significantly converted to digester gas (CO2, CH4) under anaerobic conditions (Steber and Wierich, 1986, 1987; Nowack, 1998). • With the exception of PBTC, which adsorbed poorest of all phosphonates, polyphosphonates adsorbed well on activated and digested sludge (Table 3). • Phosphonates adsorbed stronger in an acidic pH milieu than between pH 6 and 9 (Nowack, 2002b). • Water hardness improved the adsorption of phosphonates onto activated sludge (Nowack, 2002b). • With the exception of PBTC, phosphonates, once adsorbed onto activated sludge, no longer desorbed under anaerobic conditions (Reichert, 1996). • In the literature, largely high removal rates of polyphosphonates in WWTPs (N80%) are described mostly with effluent concentrations below the LOQ/LOD, except for the publication of Metzner (1990) (removal rates of 25–70%) (Table 4). However, the LODs are in such a high range (e.g., 0.057 mg/L in the case of DTPMP (Nowack, 1998)) that a contribution of phosphonates to the eutrophication of water cannot be excluded. Furthermore, no studies exist so far about the behavior of PBTC in WWTPs. • In some aerobic degradation tests, only very low adsorption rates of phosphonates onto activated sludge were observed. In particular, nitrogen-free phosphonates, such as PBTC, seem to be subject to a smaller extent of removal (Reichert, 1996).

• Already phosphonate concentrations b 1 mg/L can have a negative effect on phosphate precipitation in WWTPs. The effect can, however, be overcome by an increased precipitant concentration (Horstmann and Grohmann, 1984, 1986). • Phosphonates can cause the need of high precipitant excesses (Horstmann and Grohmann, 1984, 1986), a deficiency of phosphorus as nutrient in the biological treatment (Wenger-Oehn et al., 2005) and can lead to problems with the compliance with phosphorus limit values (König et al., 2002).

Environmental relevance of phosphonates: • Calculations (Gledhill and Feijtel, 1992; Jaworska et al., 2002) and first measurements (Schmidt et al., 2014) suggest that phosphonates are present in water in concentrations in the lower μg/L range. However, currently there are no more detailed investigations. • Phosphonates accumulate in sewage sludge in concentrations of ca. 1.2–1.5 g/kg dry substance. • The contamination of groundwater with phosphonates has been assessed to be low, but a comprehensive risk assessment is missing (Knepper, 2003). • Phosphonates are present in surface waters mainly in the form of Ca2+ and Fe2+ complexes (Nowack, 2003; Jaworska et al., 2002) as well as adsorbed to sediment particles and suspended matter (Steber and Wierich, 1986, 1987; Fischer, 1992). • Phosphonates are subject to a variety of transformation and degradation mechanisms in the aquatic environment resulting in the release of bioavailable phosphate: degradation by MnII and O2 (Nowack and Stone, 2000), hydrolysis (Schowanek and Verstraete, 1990, 1991) and metal-catalyzed photolysis (Matthijs et al., 1989; Grohmann and Horstmann, 1989). • The phosphonates investigated so far are neither mutagenic, carcinogenic nor teratogenic, and are classified as harmless for a wide range of aquatic organisms (Gledhill and Feijtel, 1992; Maise, 1984; Metzner and Nägerl, 1982; Kästner and Gode, 1983; HERA, 2004). Their degradation products AMPA and glyphosate, however, are classified as critical from a toxicological point of view (Mañas et al., 2009; de Roos et al., 2003; Eriksson et al., 2008; Hardell et al., 2011). • The contribution of phosphonates to the eutrophication of water should not be underestimated since phosphonates are subject to natural degradation mechanisms (Matthijs et al., 1989; Grohmann and Horstmann, 1989; Nowack and Stone, 2000). In some experiments, phosphonates were even stimulating the growth of green algae and cyanobacteria (Gledhill and Feijtel, 1992; Forlani et al., 2011). • Phosphonates are discussed in context with the remobilization of metals bound to sediments and activated sludge, however, the concentrations measured in surface waters so far are considered to be too low for remobilization of significant amounts of metal (Gledhill and Feijtel, 1992; Bordas and Bourg, 1998; Müller et al., 1984). 8. Conclusions Phosphonates are very stable against biological degradation. In the literature, however, predominantly good removal efficiencies on WWTPs (80–95% removal) are mentioned, which are attributed to the high tendency of phosphonates to adsorb onto sewage sludge. On the other hand, it is questionable whether phosphonates can be adequately removed in WWTPs operating with enhanced biological phosphorus removal instead of phosphate precipitation, so that there is still demand for research in this area. The LOQs of analytical methods applied for the quantification of phosphonates were relatively high. Thus, despite the removal efficiencies in WWTPs determined by these methods appear to be high, there is a risk that phosphonates contribute to a certain degree to the eutrophication of water, particularly as phosphonates undergo natural degradation such as metal-catalyzed photolysis, weak

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hydrolysis and catalytic degradation by MnII and O2. It can be assumed that at least 9000–18,600 t/a of phosphonates reach European surface waters via direct discharge and WWTPs. Phosphonate concentrations of 0.5–2.9 μg/L have been measured in some German rivers, but higher concentrations can occur locally. Moreover, at least 7300 t/a of phosphonates are discharged via sewage sludge used in agriculture in Europe. It is assumed that the concentrations of phosphonates in water bodies are currently too low for a significant heavy metal remobilizing effect by complex formation. However, if the phosphonate concentration will increase in future, remobilized metals may reach toxic concentrations for aquatic organisms or get into the drinking water via bank filtrate. Although, phosphonates are classified as harmless for a wide variety of aquatic organisms, some degradation products of aminophosphonates, such as N-(phosphonomethyl)glycine and aminomethylphosphonic acid (AMPA), have to be seen very critical from an ecotoxicological point of view. Particularly, it is strongly recommended not to discharge wastewater containing phosphonates without pretreatment, since the contribution of phosphonates to eutrophication of weak receiving waters, in particular, cannot be excluded. In the case of indirect discharges, it has to be assured that very high concentrations of phosphonates are not present in the inflow of WWTPs. Otherwise increased precipitant dosing quantities are necessary for P removal, which then might cause a deficiency of this essential nutrient in the biological treatment step. Thus, in any case there should be a pretreatment of phosphonate containing industrial partial stream wastewaters prior to discharging them into the receiving water or WWTPs. For this purpose, several physicochemical treatment methods are presented in various scientific articles. In particular, processes using metal salts or metalcontaining materials such as the precipitation/flocculation method (Rott et al., 2017a; Fettig et al., 2000; Klinger et al., 1998), PhotoFenton method (Rott et al., 2017b), UV/FeII treatment (Lesueur et al., 2005; Rott et al., 2017b) or filtration using iron-containing materials (Chen et al., 2017; Boels et al., 2010) are very promising. Whereas the use of activated carbon is not recommended for the targeted removal of phosphonates from wastewater (Klinger et al., 1998). Acknowledgements The authors are grateful for the financial support by the WillyHager-Stiftung, Stuttgart. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2017.09.223. References AAW, 1984. Internal Report 452/84806. Albright and Wilson Ltd. (Cited by OECD SIDS ATMP). ACS, 2016. Database of American Chemical Society. https://scifinder.cas.org (accessed 24.02.2016). Ammann, A.A., 2002. Determination of strong binding chelators and their metal complexes by anion-exchange chromatography and inductively coupled plasma mass spectrometry. J. Chromatogr. A 947 (2), 205–216. Bachus, H., 2003. Komplexbildner in Abbautests und in der Realität (Complexing agents in degradation tests and in reality). Textilveredlung, Melliand Textilberichte 3, 202–204. Bachus, H., Held-Beller, S., 1993. Umweltentlastende Komplexbildner mit Bestandteilen nachwachsender Rohstoffe (Complexing agents with less effects on the environment with components of renewable raw materials). Textilveredlung 28 (5), 140–146. BADSR, 1976. Bayer AG Data short Report (Cited by OECD SIDS PBTC). BADSR, 1989. Bayer AG Data short Report (Cited by OECD SIDS PBTC). Boels, L., Tervahauta, T., Witkamp, G.J., 2010. Adsorptive removal of nitrilotris(methylenephosphonic acid) antiscalant from membrane concentrates by iron-coated waste filtration sand. J. Hazard. Mater. 182 (1–3), 855–862. Bordas, F., Bourg, A.C.M., 1998. Effect of complexing agents (EDTA and ATMP) on the remobilization of heavy metals from a polluted river sediment. Aquat. Geochem. 4 (2), 201–214. Bucheli-Witschel, M., Egli, T., 2001. Environmental fate and microbial degradation of aminopolycarboxylic acids. FEMS Microbiol. Rev. 25, 69–106.

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