Nonylphenol, branched, ethoxylated

Ecotoxicological Summary

Hazard for aquatic organisms

Freshwater

Hazard assessment conclusion:

PNEC aqua (freshwater)

PNEC value:

0.8 µg/L

Assessment factor:

10

Extrapolation method:

assessment factor

PNEC freshwater (intermittent releases):

1.48 µg/L

Marine water

Hazard assessment conclusion:

PNEC aqua (marine water)

PNEC value:

0.8 µg/L

Assessment factor:

10

Extrapolation method:

assessment factor

PNEC marine water (intermittent releases):

1.48 µg/L

STP

Hazard assessment conclusion:

PNEC STP

PNEC value:

10 mg/L

Assessment factor:

10

Extrapolation method:

assessment factor

Sediment (freshwater)

Hazard assessment conclusion:

PNEC sediment (freshwater)

PNEC value:

4.6 mg/kg sediment dw

Assessment factor:

50

Extrapolation method:

assessment factor

Sediment (marine water)

Hazard assessment conclusion:

PNEC sediment (marine water)

PNEC value:

0.46 mg/kg sediment dw

Assessment factor:

500

Extrapolation method:

assessment factor

Hazard for air

Air

Hazard assessment conclusion:

no hazard identified

Hazard for terrestrial organisms

Soil

Hazard assessment conclusion:

no exposure of soil expected

Hazard for predators

Secondary poisoning

Hazard assessment conclusion:

no potential for bioaccumulation

Additional information

NP and NPEs have been widely studied for their endocrine mediated effects in the aquatic environment. The following section summarises some of the important data, as presented in a number of published reviews. Overall, the data suggest that NP has the highest endocrine activity and that effects decrease with increasing degree of ethoxylation. In the aquatic environment, the lowest concentrations showing effects (delayed larval development and larval deformities) were equivalent to 0.1 – 100 µg/L NP in a study with the Pacific oyster Crassostrea gigas by Nice HE et al.(2000). However, it was unclear whether these changes were due to endocrine disruption or other toxicity mechanisms. Most other authors reported effects only at concentrations one to several orders of magnitude higher. Regarding endocrine effects of NP on benthic organisms, not much quantitative data is available. Zulkosky AM et al. (2002) reported a 50% reduced reproduction of the estuarine species Leptocheirus plumulosus in sediment collected near the outfall of a waste water treatment plant in Jamaica Bay (New York) where the concentration of NP was determined to around 40 mg/kg dw.

EU Risk Assessment Report (RAR) (2002)

NP is cited as having estrogenic effects based on in vitro and in vivo studies. In vitro studies, for example tests with isolated hepatocytes from rainbow trout, have been used to characterize the mechanism of NP estrogenicity such as induction of vitellogenin (VTG). NP showed competitive displacement of estrogen from its receptor site (White R et al., 1994). NP has the highest estrogenic effects, and these decrease with increasing degree of ethoxylation. The estrogenic potency of NP relative to estradiol-17β is given as 0.000009 (Jobling S and Sumpter JP, 1993). Most of the available in vivo tests indicate that estrogenic effects start to occur at aquatic concentrations around 10 – 20 µg NP/L. In rainbow trout exposed to NP in a flow-through system at a measured concentration of 37 µg/L, significant reduction of the testis size and gonadosomic index (GSI) were observed. In an independent second experiment, the NOEC for VTG induction in rainbow trout was determined to be 20 µg/L whereas the NOEC for GSI and testicular growth was 54 µg/L (Jobling S et al., 1996). In a chronic test with Medaka (Oryzias latipes) starting at hatch (duration 3 months) with nominal NP concentrations of 10, 50 and 100 µg/L, testis-ova in male fish was observed as of 50 µg/L (NOEC: 10 µg/L) (Gray MA and Metcalf CD, 1997). In another long-term test with rainbow trout (Oncorhynchus mykiss) where the fish were exposed to NP for 35 days from hatch to 1, 10 or 30 µg/L (observations up to 431 days after the start of the test), Ashfield et al.(1998) reported on a significantly enhanced ovosomatic index at 30 µg/L (at this concentration body weights were also reduced) at the end of the experiment. In a 3 week static renewal test withDaphnia magna,significant reduction of reproduction was detected at 100 µg/L NP (Baldwin WS et al., 1997). NP caused deformed adults and offspring in a 30 day exposure study with Daphnia galeata mandotae already at 10 µg/L (Shurin JB and Dodson SI, 1997).

Langston et al. (2005)

The review reports on distribution and impact of estrogens and xeno-estrogens in the aquatic environment. With respect to in vitro tests, Legler J et al. (2002a) made a comparison of the estrogenic potency of different endocrine active compounds and determined the EC₅₀ value in the ER-CALUX assay for estradiol to be 0.0016 µg/L, whereas the corresponding value for NP was 57.3 µg/L (corresponding to a > 35,000 times lower potency). The same authors also concluded that the decrease of the polyethoxylate side chain leads to an increase in estrogenic potency. Legler et al. (2002b) however raised the point that caution is required regarding the extrapolation of the results obtained in vitro to the in vivo situation since aspects such as exogenous and endogenous binding and bioavailability have to be considered. Jobling Set al.(1996) arrived at the same conclusion comparing the in vitro and in vivo effects of NP. NP proved to have an up to 100 times higher estrogenic potency in fish as would have been expected fromin vitrodata. For these reasons, the European Scientific Committee for Toxicity, Ecotoxicity and the Environment (CSTEE) recommends to put the major emphasis on in vivo tests when screening chemicals for endocrine effects (CSTEE, 1999).

In the review of Langston WJ et al. (2005), in vivo data on fish mainly stem from field studies where effects found in situ were related to NP. Different authors are cited which looked at VTG induction in fish sampled from different locations, effects which could partially be correlated with NP levels found in water (Solé M et al., 2000). Besides VTG induction, abnormal gonad development (i.e. inhibition of testicular growth) was found in rainbow trout held at sites which were contaminated by alkylphenols (Harries J et al., 1997). Lye CM et al. (1999) attributed high VTG levels and testicular abnormalities to several estrogenic alkylphenols.

Although endocrine regulation in invertebrates is different from vertebrates such as fish (various forms of hermaphrodism, moulting), there are a number of studies linking the effects of substances such as alkylphenols to invertebrate development. Comber MHI et al.(1993), Baldwin WS et al.(1995) and Zou E and Fingerman S (1997) linked the presence of NP to the inhibition of molting and growth, as well as to the inhibition of testosterone metabolism in experiments with Daphnia magna. Life history effects have been observed in the copepod Tisbe battagliai at NP concentrations of 20 µg/L (Bechmann HE, 1997). Nice et al. (2000) reported on delayed larval development and larval deformities of the Pacific oyster Crassostrea gigas at NP concentrations of 0.1 – 100 µg/L. These authors also refer to altered sex ratio of C. gigas at NP concentrations ranging from 1 – 100 µg/L. However, it was not clear whether the effects were due to endocrine disruption or other toxicity mechanisms. In the case of other bivalves such as the Manila clamsTapes philippinarium, VTG levels measured in the haemolymph and the digestive tract of males increased at 100 and 200 µg/L NP (Matozzo V et al., 2003; Matozzo V and Marin MG, 2005). Studies on prosobranch snails (Marisa conuarietis) in which octylphenol (1 – 100 µg/L) and other xenostrogens such as bisphenol A were reported to cause the formation of „superfemales“ (Oehlmann J et al., 2000) could not be reproduced in more elaborated studies later on (at least for bisphenol A) (Forbes VE et al., 2007 a,b).

Coady et al. (2010)

In the review article of Coady K et al. (2010), a special chapter deals with so called secondary endpoint studies where the effects of NP on serum VTG concentrations, VTG gene transcription, estrogen receptor transcription, P450 enzyme activity and histological changes in liver, kidney and gonadal tissues are presented.

According to this summary, VTG induction (including VTG mRNA) in fish could be observed at NP concentrations ranging from 1 to 100 µg/L (Seki et al., 2003; Larsen BK et al., 2006; Van den Belt K et al.; 2004, Zha JM et al.,2007; Li MH and Wang ZR, 2005; Lerner DT et al. ,2007a; Staples C et al., 2004; Kim C et al., 2006; Zhang Z et al., 2005; Fent et al., 1999). Effects on clams in which VTG formation is, similar to vertebrate species, also under the control of the estrogen receptor, were also reported: Addition of 100 to 200 µg/L NP in case of marine clams (Matozzo V and Marin MG, 2005) or 500 µg/L NP in case of Dreisenia polymorpha (Quinn B et al., 2006) led to the induction of VTG in males. Regarding histopathology in fish tissues, changes in the GSI were reported at 10 and 500 µg/L NP by Zha JM et al. (2007) / Van den Belt K et al. (2004) and Yang FX et al. (2006). Balch G and Metcalf C (2006) and Zha JM et al. (2007) reported on the formation of ova-testis as a result of exposure to 29 and 30 µg/L NP, respectively. Previous reports were referring to 1.6 to 100 µg/L NP causing intersex and changes of the GSI (Miles-Richardson SR et al., 1999; Gray MA and Metcalf CD, 1997; Jobling S et al., 1996; Staples C et al.; 2004). With respect to changes of enzyme activities, Lerner DT et al. (2007b) observed effects such as decreased thyroid hormone levels, decreased gill sodium-potassium-activated ATPase activity (an enzyme considered to be of importance for the salt secretion and osmoregulation) in Atlantic salmon exposed to NP at 6.5 µg/L. Other enzymes which were suppressed as a result of the exposure to NP concentrations of 30 - 5000 µg/L included CYP1A, CYP3A and ethoxyresorufin-O-deethylase (Sturve J et al., 2006; Vacaro E et al., 2005). A moult-promoting steroid hormone of importance for invertebrates, 20-hydroxyectysone, was significantly decreased in mysid shrimp at 30 µg/L NP (Hirano M et al., 2009).

PNEC water

The PNEC calculations for freshwater and marine environments are based on the lowest chronic value of 7.7 µg/L (NOEC determined with NPE-1.5 for the salwater species mysid shrimp in a 28 day test). This study was selected for PNEC derivation as the test substance NPE-1.5 has a composition closely resembling that of NPEO, except that NPE-1.5 also contains 3.8% of NP.

The PNEC calculation for intermittent release was basedon the lowest available valid acute value of 148 µg/L (48 h LC₅₀determined for Daphnia magna) and an assessment factor of 100.

Regarding endocrine disruption, the EU RAR completed in 2002 concluded that „data exist indicating toxicity at lower concentrations than the concentrations at which estrogenic effects were observed. Therefore, the calculated PNEC (of 0.33 µg/L) should be protective for estrogenic effects in fish as well.“ (EU RAR, 2002).

The review of Langston WJ et al. (2005) summarized the reported data by stating that NOECs for OP and NP are of the order of a few µg/L, although general toxicity endpoints may be lower than endocrine effects. Similarly to the EU RAR of 2002, these authors considered general toxicity effects as occurring at lower concentrations than endocrine effects.

According to Coady K et al. (2010), “(…)the current U.S. chronic WQC of 6.6 µg/L for NP appears to be sufficiently protective for freshwater communities in light of the recent literature on this substance.“

Based on the above considerations, it can be concluded that the aquatic PNECs of 0.8 µg/L (freshwater and marine water) and 1.48 µg/L (intermittent release) established for NPEO can be considered as protective also for endocrine disruption endpoints.

PNEC sediment

The PNEC calculation for freshwater and marine sediment is based onthe lowest EC₁₀ value (231 mg/kg dw) from a study with C. riparius exposed to NP. An assessment factor of 50 was applied for freshwater and 500 for marine water.

The PNECs for sediment organisms of 4.6 mg/kg dw (4.6 µg/g dw) for freshwater and 0.46 mg/kg dw (0.46 µg/g dw) for marine species can reasonably be assumed to protect also for endocrine effects on sediment organisms in the light of the data generated by Zulkosky AM et al. (2002) showing 50% reduction of reproduction in Leptocheirus plumulosus at around 40 mg/kg dw NP.

Conclusion on classification

Acute LC50 values for NPEO are in the range of 0.11 to 0.716 mg/L. A NOEC of 0.006 mg/L from a study on NP was considered relevant for the long-term aquatic toxicity of NPEO. Furthermore, the substance is inherently biodegradable. Based on these data, the following environmental classification is proposed: Aquatic acute 1 - H400 (very toxic to aquatic life) and Aquatic chronic 1 - H410 (Very toxic to aquatic life with long lasting effect) according to CLP criteria (EC 1272/2008). Since the lowest acute L(E)C50 is between 0.1 and 1 mg/L, an acute M factor of 1 is applied. Since the lowest chronic NOEC is between 0.001 and 0.01 mg/L and the substance is non-readily biodegradable, a chronic M factor of 10 is applied.