Next Article in Journal
Molecular Cloning and Functional Characterization of Tibetan Porcine STING
Previous Article in Journal
Non Coding RNAs and Viruses in the Framework of the Phylogeny of the Genes, Epigenesis and Heredity
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:

Occurrence and Biodegradation of Nonylphenol in the Environment

School of Environment Science and Spatial Informatics, China University of Mining and Technology, Xuzhou 221116, China
School of Chemical Engineering and Technology, China University of Mining and Technology, Xuzhou 221116, China
Author to whom correspondence should be addressed.
Int. J. Mol. Sci. 2012, 13(1), 491-505;
Submission received: 25 November 2011 / Revised: 8 December 2011 / Accepted: 20 December 2011 / Published: 4 January 2012


Nonylphenol (NP) is an ultimate degradation product of nonylphenol polyethoxylates (NPE) that is primarily used in cleaning and industrial processes. Its widespread use has led to the wide existence of NP in various environmental matrices, such as water, sediment, air and soil. NP can be decreased by biodegradation through the action of microorganisms under aerobic or anaerobic conditions. Half-lives of biodegradation ranged from a few days to almost one hundred days. The degradation rate for NP was influenced by temperature, pH and additions of yeast extracts, surfactants, aluminum sulfate, acetate, pyruvate, lactate, manganese dioxide, ferric chloride, sodium chloride, hydrogen peroxide, heavy metals, and phthalic acid esters. Although NP is present at low concentrations in the environment, as an endocrine disruptor the risks of long-term exposure to low concentrations remain largely unknown. This paper reviews the occurrence of NP in the environment and its aerobic and anaerobic biodegradation in natural environments and sewage treatment plants, which is essential for assessing the potential risk associated with low level exposure to NP and other endocrine disruptors.

1. Introduction

Since the early 1990s, a lot of research has been performed concerning the endocrine disruptors which are widespread in the environment. Previous studies have demonstrated that nonylphenol (NP) is one of the endocrine disruptors, and many studies have shown that NP can exert adverse effects to an ecosystem.
NP is usually reacted to produce nonylphenol polyethoxylates (NPEs). NPEs are mainly used in a number of industrial processes and products, including cleaners, detergents and plastics. Annual world production of NPEs was about 520,000 tons in 1995, and the demand is increasing annually. In China, the production of NPEs is about 50,000 tons per year, and most enters into the aquatic environment [1]. The wide variety in use makes products containing NPEs potential sources of diffuse emissions of NPEs and NP. NPEs can biodegrade into NPs in sewage treatment works or in the environment. NP is persistent, lipophilic and tends to bioaccumulate more than the parent compounds [2,3]. Due to the endocrine potential of NP, the “Oslo and Paris Commission for the Protection of the Marine Environment of the north-east Atlantic” called for phasing out the use of NPEs in domestic cleaning agents by 1995 and in industrial cleaning agents by 2000. Following these recommendations, many countries, such as Sweden, Belgium, Great Britain, Germany, Holland, etc., have drastically limited the use of NPEs. Switzerland has completely banned the use of these substances [4]. In 2006, the U.S. Environmental Protection Agency (EPA) released final aquatic life ambient water quality criteria for NP, which recommends NP concentrations in both freshwater (28 μg/L, acute; 6.6 μg/L, chronic) and saltwater (7.0 μg/L, acute; 1.7 μg/L chronic). In Canada and Japan the use and production of NPEs are strictly monitored. But in many developing countries, such as China and India, no schedule was made to decrease the use of NP step by step. Meanwhile, the production of NPEs in these countries was increased annually.

2. Occurrence of Nonylphenol in the Environment

NP is a compound which has numerous isomers. The side chain has nine carbons and can be attached to phenol at different points on the ring, thus producing different isomers. 4-NP is the most common commercial forms of NP, which is often used in experimentation and the analysis [5,6]. At room temperature, NP is a pale yellow liquid with an approximate molecular weight of 215 to 220 g/mol and a specific gravity of 0.953 g/mL at 20 °C. It has a dissociation constant (pKa) of 10.7 ± 1.0 and an Octanol-Water Partition Coefficient (log Kow) between 3.8 and 4.8, and exhibits both pH- and temperature-dependent solubility, showing values of 6,350 μg/L at pH 5 and 25 °C [7].
NP is the biological breakdown products of widely used nonionic surfactant, NPEs, which is directly discharged into the environment. NP has been widely found in the environment: in water, sediment, air, soil, aquatic organisms and even human food.

2.1. Water and Sediment

2.1.1. Surface Water and Sediment

Generally, NP occurs in the aquatic environment with concentrations varying widely in surface water from tens of ng/L [8,9] to dozens of mg/L [10]. In sediment, NP concentrations were higher than that of surface water. Table 1 shows the occurrence and distribution of NP in surface water and sediment in many countries. Because of the implementation of the European Directive 2003/53/EC, NP concentrations in European countries were much lower than in Asia.
Duong C.N. et al. [11] estimated the occurrence and distribution of NP in Korea and seven other Asian countries including Laos, Cambodia, Vietnam, China, Indonesia, Thailand and Malaysia. The results showed that the NP concentrations in samples of most Asian countries were at a higher level in comparison to those reported in European countries, America and Japan.
The domestic and industrial wastewater produced, as well as surface runoff, could possibly be sources of NP in the aquatic environment. The distribution and characteristics of pollution sources along the river affected the spatial variation of NP [1]. There was a direct relationship between concentrations of NP and the presence of urban or industrial activities near the sampling point [12]. Inadequately treated domestic wastewater caused high concentrations of NP in aquatic environment. Concentrations in the Kaoping River’s polluted tributaries were higher due to inadequate wastewater treatment in these regions [13], which caused high risk downstream of the river. The concentrations of NP in the Jialu River ranged from 75.2 to 1520 ng/L. Zhengzhou city is regarded as the main discharge source to this river as the annual discharge of NP from its urban zone to the river was 726 kg [14]. All of these results demonstrated that even a small river without adequately treated domestic wastewater could cause high mass loadings of NP and high risk [13]. In addition, many other factors affect the variation of NP in surface water, such as temperature, flow rate, and biodegradation etc. Temperature was the key factor affecting the seasonal variation of NP in water and suspended particles. In Seine River Estuary, NP maximum levels in the dissolved phase and in the suspended particulate matter were observed during winter periods while significant decreases were observed during spring and autumn periods [9]. These declines could be ascribed to maximum biological activities during these seasons. However, in terms of most water and suspended particles samples in Lanzhou Reach of Yellow River, concentrations were higher in warmer seasons than in colder seasons [1]. The higher NP concentrations in the warm season confirm the relationship between NP contamination and sewage discharge into the surface water, because more detergents, showers with shower cream, and plastic ware were used in the warm season compared to the cold one [15]. As to the flow rate, concentrations of NP in the water’s low flow period were higher than in its high flow period due to a low dilution factor [13,16].
Due to its hydrophobic properties, NP in surface water tends to be absorbed by sediment particles, which caused a preferential accumulation in sediments [8]. The reported sedimentary concentrations of NP were from a few μg/kg dry weight (dw) up to several hundred mg/kg dw (Table 1). The most contaminated sediment was found at Lake Donghu, China in 2003 [15]. NP was found at the highest concentration at 119,100 μg/kg dw. Similar circumstances were observed in Wuhan urban lakes [17] and Pearl River of China [18,19]. In Pearl River of China some researchers found there is a positive correlation between NP and total organic carbon (TOC), which indicated that sedimentary organic carbon (SOC) is a key factor in controlling the distributions of the endocrine disruptors [18,19]. Nowadays even in some European areas the NP concentrations in sediment were very high. For example, in the Danube River maximum sediment concentrations were 2,830 μg/kg dw for nonylphenol [20]. Table 1 shows the NP concentrations in samples of some countries in recent years. Generally speaking, the NP concentrations in sediment were at a higher level in developing countries than in developed countries.
High NP accumulation made the sediments a long-term pollutant sink and reservoir. The adsorbed compounds can be released into the water phase and become source of contaminants when hydraulic regimes of rivers change. However, De Weert et al. [21] thought that NP in the water phase was more available for biodegradation than in sediments. When the rate of biodegradation in the water is higher than the rate of desorption, NP is generally biodegraded in the sediment-water interphase and will not reach the bulk water, which may result in a limited hazard for the organisms in the aquatic environment. The effects of changing conditions on desorption and biodegradation processes are essential for adequate prediction of the fate of pollutants and the ecological effects of polluted sediments.

2.1.2. Groundwater

Groundwater is of special interest because it makes up about twenty percent of the world’s fresh water supply and it is extraordinarily vulnerable to contamination by a variety of contaminants due to urban activities [36]. Micropollutants may enter the ground water nearly un-attenuated by bank filtration of affected surface waters or by infiltration or artificial recharge of treated wastewater into groundwater [37]. Contamination of groundwater is directly linked to the transport of the pollutant within the soil column supporting the advective and diffusional flow system, the geochemistry of the groundwater, and the overall groundwater flow [38].
Generally speaking, concentrations of NP in the groundwater were very low. In some area NP was not detected [38]. When NP was used as one of indicators for assessing anthropogenic impact on urban surface and groundwater in the cities of Halle/Saale and Leipzig (Germany), concentrations of NP were observed about 100 ng/L [39]. Loos R. et al. [37] collected and analyzed 164 individual groundwater samples from 23 European Countries. NP was found in 11% of the samples (with a LOD of 30 ng/L), with a maximum concentration of 3.8 μg/L, exceeding the European groundwater quality standard for pesticides of 0.1 μg/L for several samples. In Austria, the most abundant industrial chemicals in groundwater samples were NP, occurring in about half of the samples. The maximum concentration of NP was 1,500 ng/L, and the 90th percentile of NP was 424 ng/L [40]. The NP pollution of groundwater may threaten human and ecosystem health.

2.1.3. Drinking Water

Recently, drinking water safety has received significant attention. Contaminants, such as NP, in drinking water might pose health risks to some residents. NP were detected in bottled water with the concentration of about 7.9 ng/L [41]. Li X. et al. [42] investigated the 4-NP level in tap water in Guangzhou (China) using gas chromatography–mass spectrometry with negative chemical ionization. Five of the tap water samples from six drinking water plants were found to contain 4-NP both in June and December. The highest concentration in tap water for 4-NP was 1,987 ng/L. In Chongqing, another major city of southwestern China, the 4-NP removal rate by the water treatment process varied in a range from 62% to 94%, resulting in a considerably high residual 4-NP concentration in drinking water in July (0.1–2.7 μg/L) [43]. Although the daily intake values of 4-NP for a human is much lower than their tolerable daily intake (TDI) values, which are 5 μg/kg body weight for NP [42], more attention should be paid to the ecological risk.

2.1.4. Wastewater Treatment Plants

NP are the most abundant compounds in raw wastewater as well as in effluents from all the treatment stages of sewage treatment plants. In influent wastewater, concentrations of NP ranging from 0.08 to 96.4 μg/L [4449] have been reported by investigators. Biodegradation was the main removal pathway of NP, as it was more effective in removing NP from the aqueous phase than physical treatment [50]. A wide range of microorganisms were involved in NP biodegradation via different degradation pathways, which reduced the possible risk of NP in the environment under aerobic conditions [51]. Removal rates of NP ranging from 13.6 to >99% have been reported in literature [50,5254]. Generally, the elimination efficiency varied between 73% and 92% [46]. In the European Union Wastewater Treatment Plants (WWTPs) 4-NP has shown remarkable decreasing influent and effluent concentrations since the implementation of Directive 2003/53/EC [46].
In a wastewater treatment plant many factors can influence the removal of NP, such as influent load, water quality of influents, plant configurations, hydraulic residence time (HRT), sludge retention time (SRT), biomass characteristics and the environmental conditions [50,52,53,55,56]. Longer HRT or SRT and greater microbial activity appear to have a positive influence on the ability of the activated sludge system to eliminate NP [50,57]. At aerobic conditions the NP degradation potential was affected by changes in pH value or temperature, and by the addition of yeast extract, aluminum sulfate, hydrogen peroxide or surfactant [58]. Nie et al. thought that warm temperature and a high MLSS concentration would benefit the removal of NP from wastewater in the Anaerobic/Anoxic/Oxic (A/A/O) bioreactor [59].
During wastewater treatment, NP accumulated in sewage sludge at a concentration of several hundred mg/kg [48,60,61]. Because NP may be highly toxic to organisms, further research is needed to evaluate their degradability and set up disposal procedures to minimize the environmental impact produced by the use of sludge for agricultural purposes [48].

2.2. Air

The occurrence of NP in the air was ubiquitous in urban, remote, industrialized, coastal regions, and even in the central part of the North Sea (Table 2). NP is not produced naturally so their presence in the environment is the consequence of anthropogenic activity [62]. As NP present short atmospheric life they are not considered as environmentally persistent or subjected to significant long-range transport. Wind direction did not show significant influence on NP concentration. Therefore the atmospheric NP originated from local sources [63]. For this reason, the concentration of NP was higher in a densely populated and more polluted urban area [64] than remote area [65]. As a semivolatile organic compound, NP can vaporize into atmosphere from wastewater discharges, wastewater treatment plant effluents (liquid and sludge) or polluted surface waters [62,63]. The NP occurrence in the atmosphere may be an important human and ecosystem health issue in the world [64].
The atmospheric concentration of NP showed declining trends from land to the open sea, suggesting that the atmosphere is a significant pathway for the transport of alkylphenols in the environment [67]. Sea could be an important sink for the NP, and might be as a potential source for the occurrences of NP in the oceans and remote area [68]. In the winter, atmospheric deposition was dominant. However in the warm seasons, re-volatilization might happen [68]. The similar seasonal trends were observed in the lower Hudson River Estuary [66]. At all the sampling sites, gas-phase NP concentrations were significantly higher during the summer (June–September) than during the fall and early winter (October–December). Temperature might be a critical factor contributing to the seasonal trends of NP in atmosphere.

2.3. Soil

NP can be introduced into soils in various ways, for example from atmospheric deposition, from soil amendment with sewage sludge, and from wastewater for irrigation of agricultural land [69]. During wastewater treatment, large quantities of NP can be quickly sorbed by the organic rich solid phase and eventually concentrated in biosolids with levels from a few mg/kg up to several thousand mg/kg [70,71]. Biosolids are often used as fertilizer on agricultural soils and additionally it improves soil structure and aids the recycling of nutrients and organic matter [72]. NP could be introduced to soils via land application of biosolids, potentially leading to accumulation in soils and crops. NP was detected at μg/kg levels in the biosolids-amended soils [73] and the wastewater irrigated soils [74].
Although NP is capable of being leached from soil, its short half-life means that its passage from soil to freshwater will be low [75]. Previous studies showed that no NP will accumulate over time and plant uptake or water quality impairment will be minimal [7577]. This suggests that NP in the soils most probably pose very low risks to the soil and freshwater ecosystems and even human health [74,75].

3. Biodegradation of Nonylphenol in the Environment

NP can be decreased by biodegradation in the water, sediment and soils through the action of microorganisms. In the Jialu River, about 23.7% of total decrease in NP concentration was caused by biodegradation [14]. As there are many chemical and environmental factors which influence biodegradation of NP, half-lives (t1/2) for NP aerobic degradation in sewage sludge and sediments ranged from 1.1 to 99.0 days [7880]. Aerobic degradation rate for NP was enhanced by shaking, increased temperature and the addition of yeast extract (5 mg/L) and surfactants such as brij 30 or brij 35 [7880]. The addition of aluminum sulfate, hydrogen peroxide, Pb, Cd, Cu, Zn, phthalic acid esters (PAEs), and NaCl inhibited NP degradation [78,80]. Reduced levels of ammonium, phosphate, and sulfate also delayed the aerobic degradation rate for NP [79]. And the optimal pH value for NP biodegradation was 7.0 [78]. Biodegradation ability was also related to light intensity in some microorganisms, such as C. vulgaris [81].
NP degradation under anaerobic conditions has only recently been demonstrated. Half-lives of anaerobic degradation ranged from 23.9 to 69.3 days [82,83]. Anaerobic degradation rate for NP was enhanced by increasing temperature and the addition of yeast extract or surfactants such as brij 30 or brij 35. The addition of aluminum sulfate, acetate, pyruvate, lactate, manganese dioxide, ferric chloride, sodium chloride, heavy metals, and phthalic acid esters inhibit the degradation rate. The high-to-low order of degradation rates was: sulfate-reducing conditions > methanogenic conditions > nitrate-reducing conditions [82,83].
Upon entering soil, NP can undergo a number of reactions (e.g., sorption, biodegradation, leaching, plant uptake) that ultimately control its fate and potential environmental hazard [75]. The concentrations of NP sharply declined with increasing soil depth, indicating limited soil leaching of this compound [73]. Previous studies showed that different NP isomers exhibited different degradation rates, but only minimal amounts of all isomers persisted after 45 d [76,84]. Das and Xia found that isomers with a-methyl-a-propyl structure transformed significantly slower than those with less branched tertiary a-carbon and those with secondary a-carbon [70]. The main factors controlling the degradation and the rate of degradation are initial concentration of NP, soil parameters, environmental conditions, as well as agricultural practices [84,85]. Trocme et al. showed that NP degraded rapidly during incubation at low concentrations, but was more persistent at higher concentrations [86]. Appropriate pH value, temperature and the aeration status in soils contribute to increase microbial activity, and consequently enhance NP degradation. The addition of different substrates such as yeast extract, compost, or brij 35 also changed the microbial community and thus affected NP degradation in soil [87]. Some treatments, such as sludge centrifugation and lime stabilization, decrease the rate of mineralization and significantly lower degradation [85].
During recent years, a lot of research has been performed concerning the molecular mechanism for degradation of nonylphenol by a number of different strains. Understanding in more detail the molecular events in degradation of nonylphenol were illuminated from two different pathways. The ring cleavage pathway was elucidated through gene cloning and biochemical studies [88,89], which involves two critical steps. First, the aromatic ring was monohydroxylated by a multicomponent phenol hydroxylase at the ortho position. Then aromatic ring was cleaved either by catechol 1,2-dioxygenase (C12O) (which is responsible for the ortho-pathway) or catechol 2,3-dioxygenase (C23O) (which is responsible for the meta-pathway). Zhang et al. [88] investigated the changes of possible key catabolic genes during the degradation of NP in natural water microcosms and found that the copy number of catechol 2,3-dioxygenase (C23O) DNA increased significantly during NP degradation. This result suggested that meta-cleaving pathway might be involved in the degradation of NP natural water microcosms. However, Nguyen et al. [89] found that most of the isolated alkylphenol-degrading bacteria are able to degrade long-chain alkylphenols via multicomponent phenol hydroxylase and the ortho-cleavage pathway. The ipso-hydroxylation pathway was responsible for the removal of the alkyl chain from NP by Sphingomonas strains [9096], in which NP isomers were initially hydroxylated at the ipso-position forming dienones, and subsequently the nonyl chain shifts to the oxygen atom in the introduced hydroxyl group to form alkoxyphenols, from which the alkyl moieties can be easily detached as alcohols by known mechanisms [95,97].

4. Conclusions

NP is a virtually ubiquitous contaminant in the environment. The occurrence of NP has been reported around the world in waters, sediment, airs and soils. It can be decreased by biodegradation in natural environments and sewage treatment plants through the action of microorganisms under aerobic or anaerobic conditions. Although NP is present at low concentrations, the risks of long-term exposure to low concentrations remain largely unknown. More research needs to be done to determine the potential human and environmental health risks posed by exposure to NP in the environment.


  1. Xu, J.; Wang, P.; Guo, W.F.; Dong, J.X.; Wang, L.; Dai, S.G. Seasonal and spatial distribution of nonylphenol in Lanzhou Reach of Yellow River in China. Chemosphere 2006, 65, 1445–1451. [Google Scholar]
  2. Ekelund, R.; Granmo, A.; Magnusson, K.; Berggren, M. Biodegradation of 4-nonylphenol in seawater and sediment. Environ. Pollut 1993, 79, 59–61. [Google Scholar]
  3. Ahel, M.; Giger, W.; Koch, M. Behavior of Alkylphenol polyethoxylate surfactants in the aquatic environment. 1. Occurrence and transformation in sewage-treatment. Water Res 1994, 28, 1131–1142. [Google Scholar]
  4. Ferrara, F.; Fabietti, F.; Delise, M.; Bocca, A.P.; Funari, E. Alkylphenolic compounds in edible molluscs of the Adriatic Sea (Italy). Environ. Sci. Technol 2001, 35, 3109–3112. [Google Scholar]
  5. Giger, W.; Brunner, P.H.; Schaffner, C. 4-nonylphenol in sewage-sludge—Accumulation of toxic metabolites from nonionic surfactants. Science 1984, 225, 623–625. [Google Scholar]
  6. Sekela, M.; Brewer, R.; Moyle, G.; Tuominen, T. Occurrence of an environmental estrogen (4-nonylphenol) in sewage treatment plant effluent and the aquatic receiving environment. Water Sci. Technol 1999, 39, 217–220. [Google Scholar]
  7. Ahel, M.; Giger, W. Aqueous solubility of alkylphenols and alkylphenol polyethoxylates. Chemosphere 1993, 26, 1461–1470. [Google Scholar]
  8. Brix, R.; Postigo, C.; Gonzalez, S.; Villagrasa, M.; Navarro, A.; Kuster, M.; de Alda, M.J.L.; Barcelo, D. Analysis and occurrence of alkylphenolic compounds and estrogens in a European river basin and an evaluation of their importance as priority pollutants. Anal. Bioanal. Chem 2010, 396, 1301–1309. [Google Scholar]
  9. Cailleaud, K.; Forget-Leray, J.; Souissi, S.; Lardy, S.; Augagneur, S.; Budzinski, H. Seasonal variation of hydrophobic organic contaminant concentrations in the water-column of the Seine Estuary and their transfer to a planktonic species Eurytemora affinis (Calanoid, copepod). Part 2: Alkylphenol-polyethoxylates. Chemosphere 2007, 70, 281–287. [Google Scholar]
  10. Peng, X.Z.; Yu, Y.J.; Tang, C.M.; Tan, J.H.; Huang, Q.X.; Wang, Z.D. Occurrence of steroid estrogens, endocrine-disrupting phenols, and acid pharmaceutical residues in urban riverine water of the Pearl River Delta, South China. Sci. Total Environ 2008, 397, 158–166. [Google Scholar]
  11. Duong, C.N.; Ra, J.S.; Cho, J.; Kim, S.D.; Choi, H.K.; Park, J.H.; Kim, K.W.; Inam, E.; Kim, S.D. Estrogenic chemicals and estrogenicity in river waters of South Korea and seven Asian countries. Chemosphere 2010, 78, 286–293. [Google Scholar]
  12. Vitali, M.; Ensabella, F.; Stella, D.; Guidotti, M. Nonylphenols in freshwaters of the hydrologic system of an Italian district: Association with human activities and evaluation of human exposure. Chemosphere 2004, 57, 1637–1647. [Google Scholar]
  13. Chen, T.C.; Yeh, Y.L. Ecological risk, mass loading, and occurrence of nonylphenol (NP), NP mono-, and diethoxylate in Kaoping River and its tributaries, Taiwan. Water Air Soil Poll 2010, 208, 209–220. [Google Scholar]
  14. Zhang, Y.Z.; Tang, C.Y.; Song, X.F.; Li, F.D. Behavior and fate of alkylphenols in surface water of the Jialu River, Henan Province, China. Chemosphere 2009, 77, 559–565. [Google Scholar]
  15. Xue, X.; Wu, F.; Zhang, X.; Deng, N. Occurrence of endocrine disrupting compounds in rivers and lakes of Wuhan City, China. Fresenius Environ. Bull 2008, 17, 203–210. [Google Scholar]
  16. Shue, M.F.; Chen, F.A.; Chen, T.C. Total estrogenic activity and nonylphenol concentration in the Donggang River, Taiwan. Environ. Monit. Assess 2010, 168, 91–101. [Google Scholar]
  17. Wu, Z.B.; Zhang, Z.; Chen, S.P.; He, F.; Fu, G.P.; Liang, W. Nonylphenol and octylphenol in urban eutrophic lakes of the subtropical China. Fresenius Environ. Bull 2007, 16, 227–234. [Google Scholar]
  18. Gong, J.; Chen, D.Y.; Yang, Y. Occurrence of endocrine-disrupting chemicals in riverine sediments from the Pearl River Delta, China. Mar. Pollut. Bull 2011, 63, 556–563. [Google Scholar]
  19. Gong, J.; Yang, Y.; Chen, D.Y.; Ran, Y. Sequential ASE extraction of alkylphenols from sediments: Occurrence and environmental implications. J. Hazard. Mater 2011, 192, 643–650. [Google Scholar]
  20. Micic, V.; Hofmann, T. Occurrence and behaviour of selected hydrophobic alkylphenolic compounds in the Danube River. Environ. Pollut 2009, 157, 2759–2768. [Google Scholar]
  21. De Weert, J.; Vinas, M.; Grotenhuis, T.; Rijnaarts, H.; Langenhoff, A. Aerobic nonylphenol degradation and nitro-nonylphenol formation by microbial cultures from sediments. Appl. Microbiol. Biotechnol 2010, 86, 761–771. [Google Scholar]
  22. Arditsoglou, A.; Voutsa, D. Determination of phenolic and steroid endocrine disrupting compounds in environmental matrices. Environ. Sci. Pollut. Res. Int 2008, 15, 228–236. [Google Scholar]
  23. Jonkers, N.; Kohler, H.P.E.; Dammshauser, A.; Giger, W. Mass flows of endocrine disruptors in the Glatt River during varying weather conditions. Environ. Sci. Pollut. Res. Int 2009, 157, 714–723. [Google Scholar]
  24. Loos, R.; Locoro, G.; Contini, S. Occurrence of polar organic contaminants in the dissolved water phase of the Danube River and its major tributaries using SPE-LC-MS2 analysis. Water Res 2010, 44, 2325–2335. [Google Scholar]
  25. Jonkers, N.; Sousa, A.; Galante-Oliveira, S.; Barroso, C.M.; Kohler, H.P.E.; Giger, W. Occurrence and sources of selected phenolic endocrine disruptors in Ria de Aveiro, Portugal. Environ. Sci. Pollut. Res. Int 2010, 17, 834–843. [Google Scholar]
  26. Mayer, T.; Bennie, D.; Rosa, F.; Rekas, G.; Palabrica, V.; Schachtschneider, J. Occurrence of alkylphenolic substances in a Great Lakes coastal marsh, Cootes Paradise, ON, Canada. Environ. Pollut 2007, 147, 683–690. [Google Scholar]
  27. Arditsoglou, A.; Voutsa, D. Partitioning of endocrine disrupting compounds in inland waters and wastewaters discharged into the coastal area of Thessaloniki, Northern Greece. Environ. Sci. Pollut. Res. Int 2010, 17, 529–538. [Google Scholar]
  28. Quednow, K.; Puttmann, W. Temporal concentration changes of DEET, TCEP, terbutryn, and nonylphenols in freshwater streams of Hesse, Germany: Possible influence of mandatory regulations and voluntary environmental agreements. Environ. Sci. Pollut. Res. Int 2009, 16, 630–640. [Google Scholar]
  29. Rice, C.P.; Isabelle, S.; Loyo-Rosales, J.E.; Edward, L.; Roger, T.; Laura, F. Alkylphenol and alkylphenol-ethoxylates in carp, water, and sediment from the Cuyahoga River, Ohio. Environ. Sci. Technol 2003, 37, 3747–3754. [Google Scholar]
  30. Writer, J.H.; Brown, G.K.; Taylor, H.E.; Kiesling, R.L. Anthropogenic tracers, endocrine disrupting chemicals, and endocrine disruption in Minnesota lakes. Sci. Total Environ 2010, 409, 100–111. [Google Scholar]
  31. Voutsa, D.; Schaffner, C.; Giger, W. Benzotriazoles, alkylphenols and bisphenol a in municipal wastewaters and in the Glatt River, Switzerland. Environ. Sci. Pollut. Res. Int 2006, 13, 333–341. [Google Scholar]
  32. Lara-Martin, P.A.; Gomez-Parra, A.; Barcelo, D.; Gonzalez-Mazo, E. Presence of surfactants and their degradation intermediates in sediment cores and grabs from the Cadiz Bay area. Environ. Pollut 2006, 144, 483–491. [Google Scholar]
  33. Wang, B.; Zhao, S.; Wang, Y.; Yu, F. Analysis of six phenolic endocrine disrupting chemicals in surface water and sediment. Chromatographia 2011, 74, 297–306. [Google Scholar]
  34. Grund, S.; Schoenenberger, R.; Suter, M.J.F.; Giesy, J.P. The endocrine disrupting potential of sediments from the Upper Danube River (Germany) as revealed by in vitro bioassays and chemical analysis. Environ. Sci. Pollut. Res. Int 2011, 18, 446–460. [Google Scholar]
  35. Pojana, G.; Jonkers, N.; Marcomini, A. Natural and synthetic endocrine disrupting compounds (EDCs) in water, sediment and biota of a coastal lagoon. Environ. Int 2007, 33, 929–936. [Google Scholar]
  36. Tubau, I.; Vazquez-Sune, E.; Carrera, J.; Gonzalez, S.; Petrovic, M.; de Alda, M.J.L.; Barcelo, D. Occurrence and fate of alkylphenol polyethoxylate degradation products and linear alkylbenzene sulfonate surfactants in urban ground water: Barcelona case study. J. Hydrol 2010, 383, 102–110. [Google Scholar]
  37. Loos, R.; Locoro, G.; Comero, S.; Contini, S.; Schwesig, D.; Werres, F.; Balsaa, P.; Gans, O.; Weiss, S.; Blaha, L.; et al. Pan-European survey on the occurrence of selected polar organic persistent pollutants in ground water. Water Res 2010, 44, 4115–4126. [Google Scholar]
  38. Hildebrandt, A.; Lacorte, S.; Barcelo, D. Assessment of priority pesticides, degradation products, and pesticide adjuvants in groundwaters and top soils from agricultural areas of the Ebro river basin. Anal. Bioanal. Chem 2007, 387, 1459–1468. [Google Scholar]
  39. Strauch, G.; Moder, M.; Wennrich, R.; Osenbruck, K.; Glaser, H.R.; Schladitz, T.; Muller, C.; Schirmer, K.; Reinstorf, F.; Schirmer, M. Indicators for assessing anthropogenic impact on urban surface and groundwater. J. Soil Sediment 2008, 8, 23–33. [Google Scholar]
  40. Hohenblum, P.; Gans, O.; Moche, W.; Scharf, S.; Lorbeer, G. Monitoring of selected estrogenic hormones and industrial chemicals in groundwaters and surface waters in Austria. Sci. Total Environ 2004, 333, 185–193. [Google Scholar]
  41. Amiridou, D. Alkylphenols and phthalates in bottled waters. J. Hazard. Mater 2011, 185, 281–286. [Google Scholar]
  42. Li, X.; Ying, G.G.; Su, H.C.; Yang, X.B.; Wang, L. Simultaneous determination and assessment of 4-nonylphenol, bisphenol A and triclosan in tap water, bottled water and baby bottles. Environ. Int 2010, 36, 557–562. [Google Scholar]
  43. Shao, B. Nonylphenol and nonylphenol ethoxylates in river water, drinking water, and fish tissues in the area of Chongqing, China. Arch. Environ. Contam. Toxicol 2005, 48, 467–473. [Google Scholar]
  44. Nakada, N.; Tanishima, T.; Shinohara, H.; Kiri, K.; Takada, H. Pharmaceutical chemicals and endocrine disrupters in municipal wastewater in Tokyo and their removal during activated sludge treatment. Water Res 2006, 40, 3297–3303. [Google Scholar]
  45. Stasinakis, A.S.; Gatidou, G.; Mamais, D.; Thomaidis, N.S.; Lekkas, T.D. Occurrence and fate of endocrine disrupters in Greek sewage treatment plants. Water Res 2008, 42, 1796–1804. [Google Scholar]
  46. Hohne, C.; Puttmann, W. Occurrence and temporal variations of the xenoestrogens bisphenol A, 4-tert-octylphenol, and tech. 4-nonylphenol in two German wastewater treatment plants. Environ. Sci. Pollut. Res. Int 2008, 15, 405–416. [Google Scholar]
  47. Klecka, G.M. Occurrence of nonylphenol ethoxylates and their metabolites in municipal wastewater treatment plants and receiving waters. Water Environ. Res 2010, 82, 447–454. [Google Scholar]
  48. Cespedes, R.; Lacorte, S.; Ginebreda, A. Barceloa Occurrence and fate of alkylphenols and alkylphenol ethoxylates in sewage treatment plants and impact on receiving waters along the Ter River (Catalonia, NE Spain). Environ. Pollut 2008, 153, 384–392. [Google Scholar]
  49. Fernandez, M.P.; Buchanan, I. An assessment of estrogenic organic contaminants in Canadian wastewaters. Sci. Total Environ 2007, 373, 250–269. [Google Scholar]
  50. Zhou, H. Behaviour of selected endocrine-disrupting chemicals in three sewage treatment plants of Beijing, China. Environ. Monit. Assess 2010, 161, 107–121. [Google Scholar]
  51. De Weert, J. Aerobic nonylphenol degradation and nitro-nonylphenol formation by microbial cultures from sediments. Appl. Microbiol. Biotechnol 2010, 86, 761–771. [Google Scholar]
  52. Pothitou, P.; Voutsa, D. Endocrine disrupting compounds in municipal and industrial wastewater treatment plants in Northern Greece. Chemosphere 2008, 73, 1716–1723. [Google Scholar]
  53. Tan, B.L.L. Comprehensive study of endocrine disrupting compounds using grab and passive sampling at selected wastewater treatment plants in South East Queensland, Australia. Environ. Int 2007, 33, 654–669. [Google Scholar]
  54. Vogelsang, C.; Jantsch, T.G.; Tollefsen, K.E.; Liltved, H. Occurrence and removal of selected organic micropollutants at mechanical, chemical and advanced wastewater treatment plants in Norway. Water Res 2006, 40, 3559–3570. [Google Scholar]
  55. Cirja, M. Factors affecting the removal of organic micropollutants from wastewater in conventional treatment plants (CTP) and membrane bioreactors (MBR). Rev. Environ. Sci. Biotech 2008, 7, 61–78. [Google Scholar]
  56. Ying, G.G. Environmental fate of alkylphenols and alkylphenol ethoxylates—A review. Environ. Int 2002, 28, 215–226. [Google Scholar]
  57. Johnson, A.C. Comparing steroid estrogen, and nonylphenol content across a range of European sewage plants with different treatment and management practices. Water Res 2005, 39, 47–58. [Google Scholar]
  58. Chang, B.V.; Chiang, F.; Yuan, S.Y. Anaerobic degradation of nonylphenol in sludge. Chemosphere 2005, 59, 1415–1420. [Google Scholar]
  59. Nie, Y. Fate and seasonal variation of endocrine-disrupting chemicals in a sewage treatment plant with A/A/O process. Sep. Purif. Technol 2011, 84, 9–15. [Google Scholar]
  60. Gonzalez, M.M.; Martin, J.; Santos, J.L.; Aparicio, I.; Alonso, E. Occurrence and risk assessment of nonylphenol and nonylphenol ethoxylates in sewage sludge from different conventional treatment processes. Sci. Total Environ 2010, 408, 563–570. [Google Scholar]
  61. Lin, J.G.; Arunkumar, R.; Liu, C.H. Efficiency of supercritical fluid extraction for determining 4-nonylphenol in municipal sewage sludge. J. Chromatogr. A 1999, 840, 71–79. [Google Scholar]
  62. Cincinelli, A.; Mandorlo, S.; Dickhut, R.M.; Lepri, L. Particulate organic compounds in the atmosphere surrounding an industrialised area of Prato (Italy). Atmos. Environ 2003, 37, 3125–3133. [Google Scholar]
  63. Salapasidou, M.; Samara, C.; Voutsa, D. Endocrine disrupting compounds in the atmosphere of the urban area of Thessaloniki, Greece. Atmos. Environ 2011, 45, 3720–3729. [Google Scholar]
  64. Dachs, J.; van Ry, D.A.; Eisenreich, S.J. Occurrence of Estrogenic Nonylphenols in the Urban and Coastal Atmosphere of the Lower Hudson River Estuary. Environ. Sci. Technol 1999, 33, 2676–2679. [Google Scholar]
  65. Berkner, S.; Streck, G.; Herrmann, R. Development and validation of a method for determination of trace levels of alkylphenols and bisphenol A in atmospheric samples. Chemosphere 2004, 54, 575–584. [Google Scholar]
  66. Van Ry, D.A.; Dachs, J.; Gigliotti, C.L.; Brunciak, P.A.; Nelson, E.D.; Eisenreich, S.J. Atmospheric seasonal trends and environmental fate of alkylphenols in the lower Hudson River estuary. Environ. Sci. Technol 2000, 34, 2410–2417. [Google Scholar]
  67. Xie, Z.; Selzer, J.; Ebinghaus, R.; Caba, A.; Ruck, W. Development and validation of a method for the determination of trace alkylphenols and phthalates in the atmosphere. Anal. Chim. Acta 2006, 565, 198–207. [Google Scholar]
  68. Xie, Z.; Lakaschus, S.; Ebinghaus, R.; Caba, A.; Ruck, W. Atmospheric concentrations and air-sea exchanges of nonylphenol, tertiary octylphenol and nonylphenol monoethoxylate in the North Sea. Environ. Pollut 2006, 142, 170–180. [Google Scholar]
  69. Gibson, R.; Duran-Alvarez, J.C.; Estrada, K.L.; Chavez, A.; Jimenez Cisneros, B. Accumulation and leaching potential of some pharmaceuticals and potential endocrine disruptors in soils irrigated with wastewater in the Tula Valley, Mexico. Chemosphere 2010, 81, 1437–1445. [Google Scholar]
  70. Das, K.C.; Xia, K. Transformation of 4-nonylphenol isomers during biosolids composting. Chemosphere 2008, 70, 761–768. [Google Scholar]
  71. La Guardia, M.J.; Hale, R.C.; Harvey, E.; Mainor, T.M. Alkylphenol ethoxylate degradation products in land-applied sewage sludge (biosolids). Environ. Sci. Technol 2001, 35, 4798–4804. [Google Scholar]
  72. Zhang, H.; Spiteller, M.; Guenther, K.; Boehmler, G.; Zuehlke, S. Degradation of a chiral nonylphenol isomer in two agricultural soils. Environ. Pollut 2009, 157, 1904–1910. [Google Scholar]
  73. Xia, K.; Hundal, L.S.; Kumar, K.; Armbrust, K.; Cox, A.E.; Granato, T.C. Triclocarban, triclosan, polybrominated diphenyl ethers, and 4-nonylphenol in biosolids and in soil receiving 33-year biosolids application. Environ. Toxicol. Chem 2010, 29, 597–605. [Google Scholar]
  74. Chen, F.; Ying, G.G.; Kong, L.X.; Wang, L.; Zhao, J.L.; Zhou, L.J.; Zhang, L.J. Distribution and accumulation of endocrine-disrupting chemicals and pharmaceuticals in wastewater irrigated soils in Hebei, China. Environ. Pollut 2011, 159, 1490–1498. [Google Scholar]
  75. Roberts, P.; Roberts, J.P.; Jones, D.L. Behaviour of the endocrine disrupting chemical nonylphenol in soil: Assessing the risk associated with spreading contaminated waste to land. Soil Biol. Biochem 2006, 38, 1812–1822. [Google Scholar]
  76. Brown, S.; vin-Clarke, D.; Doubrava, M.; O’Connor, G. Fate of 4-nonylphenol in a biosolids amended soil. Chemosphere 2009, 75, 549–554. [Google Scholar]
  77. Petersen, S.O.; Henriksen, K.; Mortensen, G.K.; Krogh, P.H.; Brandt, K.K.; Sorensen, J.; Madsen, T.; Petersen, J.; Gron, C. Recycling of sewage sludge and household compost to arable land: Fate and effects of organic contaminants, and impact on soil fertility. Soil Till. Res 2003, 72, 139–152. [Google Scholar]
  78. Yuan, S.Y.; Yu, C.H.; Chang, B.V. Biodegradation of nonylphenol in river sediment. Environ. Pollut 2004, 127, 425–430. [Google Scholar]
  79. Chang, B.V.; Chiang, F.; Yuan, S.Y. Biodegradation of nonylphenol in sewage sludge. Chemosphere 2005, 60, 1652–1659. [Google Scholar]
  80. Chang, B.V.; Liu, C.L.; Yuan, S.Y.; Cheng, C.Y.; Ding, W.H. Biodegradation of nonylphenol in mangrove sediment. Int. Biodeterior. Biodegrad 2008, 61, 325–330. [Google Scholar]
  81. Gao, Q.T.; Wong, Y.S.; Tam, N.F.Y. Removal and biodegradation of nonylphenol by different Chlorella species. Mar. Pollut. Bull 2011, 63, 445–451. [Google Scholar]
  82. Chang, B.V.; Chiang, F.; Yuan, S.Y. Anaerobic degradation of nonylphenol in sludge. Chemosphere 2005, 59, 1415–1420. [Google Scholar]
  83. Chang, B.V.; Yu, C.H.; Yuan, S.Y. Degradation of nonylphenol by anaerobic microorganisms from river sediment. Chemosphere 2004, 55, 493–500. [Google Scholar]
  84. Sjostrom, A.E.; Collins, C.D.; Smith, S.R.; Shaw, G. Degradation and plant uptake of nonylphenol (NP) and nonylphenol-12-ethoxylate (NP12EO) in four contrasting agricultural soils. Environ. Pollut 2008, 156, 1284–1289. [Google Scholar]
  85. Kouloumbos, V.N.; Scheffer, A.; Corvini, P.F.X. Impact of sewage sludge conditioning and dewatering on the fate of nonylphenol in sludge-amended soils. Water Res 2008, 42, 3941–3951. [Google Scholar]
  86. Trocme, M.; Tarradellas, J.; Vedy, J.C. Biotoxicity and persistence of nonylphenol during incubation in a compost-sandstone mixture. Biol. Fert. Soils 1988, 5, 299–303. [Google Scholar]
  87. Chang, B.V.; Chiang, B.W.; Yuan, S.Y. Biodegradation of nonylphenol in soil. Chemosphere 2007, 66, 1857–1862. [Google Scholar]
  88. Zhang, Y.; Sei, K.; Toyama, T.; Ike, M.; Zhang, J.; Yang, M.; Kamagata, Y. Changes of catabolic genes and microbial community structures during biodegradation of nonylphenol ethoxylates and nonylphenol in natural water microcosms. Biochem. Eng. J 2008, 39, 288–296. [Google Scholar]
  89. Nguyen, N.T.; Lin, Y.W.; Huang, S.L. Analysis of bacterial degradation pathways for long-chain alkylphenols involving phenol hydroxylase, alkylphenol monooxygenase and catechol dioxygenase genes. Bioresour. Technol 2011, 102, 4232–4240. [Google Scholar]
  90. Corvini, P.F.X.; Meesters, R.J.W.; Schaffer, A.; Schroder, H.F.; Vinken, R.; Hollender, J. Degradation of a nonylphenol single isomer by Sphingomonas sp strain TTNP3 leads to a hydroxylation-induced migration product. Appl. Environ. Microbiol 2004, 70, 6897–6900. [Google Scholar]
  91. Corvini, P.F.X.; Vinken, R.; Hommes, G.; Mundt, M.; Hollender, J.; Meesters, R.; Schroder, H.F.; Schmidt, B. Microbial degradation of a single branched isomer of nonylphenol by Sphingomonas TTNP3. Water Sci. Technol 2004, 50, 189–194. [Google Scholar]
  92. Corvini, P.F.X.; Schaffer, A.; Schlosser, D. Microbial degradation of nonylphenol and other alkylphenols—Our evolving view. Appl. Microbiol. Biotechnol 2006, 72, 223–243. [Google Scholar]
  93. Corvini, P.F.X.; Hollender, J.; Ji, R.; Schumacher, S.; Prell, J.; Hommes, G.; Priefer, U.; Vinken, R.; Schaffer, A. The degradation of alpha-quaternary nonylphenol isomers by Sphingomonas sp strain TTNP3 involves a type II ipso-substitution mechanism. Appl. Microbiol. Biotechnol 2006, 70, 114–122. [Google Scholar]
  94. Corvini, P.F.X.; Mundt, M.; Schaeffer, A.; Schmidt, B. Contribution to the detection and identification of oxidation metabolites of nonylphenol in Sphingomonas sp Strain TTNP3. Biogegradation 2007, 18, 233–245. [Google Scholar]
  95. Gabriel, F.L.P.; Rentsch, D.; Giger, W.; Guenther, K. A novel metabolic pathway for degradation of 4-nonylphenol environmental contaminants by Sphingomonas xenophaga Bayram: ipso-Hydroxylation and intramolecular rearrangement. J. Biol. Chem 2005, 280, 15526–15533. [Google Scholar]
  96. Kohler, H.P.; Giger, W. ipso-Substitution—A novel pathway for microbial metabolism of endocrine-disrupting 4-nonylphenols, 4-alkoxyphenols, and bisphenol A. CHIMIA Int. J. Chem 2008, 62, 358–363. [Google Scholar]
  97. Staples, C.A.; Naylor, C.G.; Losey, B.S. C8- and C9-alkylphenols and ethoxylates: I. Identity, physical characterization, and biodegradation pathways analysis. Hum. Ecol. Risk Assess 2008, 14, 1007–1024. [Google Scholar]
Table 1. Nonylphenol (NP) levels in surface water and sediment samples.
Table 1. Nonylphenol (NP) levels in surface water and sediment samples.
NP levels

Surface water (μg/L)Sediment (μg/kg)LocationDetected timeReference
0.034–0.59938.4–863.0Lanzhou Reach of Yellow River, ChinaJuly and November 2004[1]
NA-0.53LOD-79Llobregat basin, Spain2005–2006[8]
0.112266Thermaiko Gulf, Greece[22]
0.227-Loudias River, Greece[22]
<LOD-310-Kaoping River and its tributaries, TaiwanJuly 2004–December 2005[13]
0.075–1.520-Jialu River, China September2007[14]
0.266 ± 0.028-Yeongsan and Seomjin rivers, Southern Korea2008[11]
0.043 ± 0.005-Ton River in Souan Mone, Pear Lart and Park Ton, Laos2008[11]
< LOD-Siem Reap River, Chong Srok area, Cambodia2008[11]
2.097 ± 0.212-Long Xuyen city and nearby area, Vietnam2008[11]
0.372 ± 0.040-Fenhe River, China2008[11]
0.039 ± 0.005-Cikamasan, Cisarua, Indonesia2008[11]
0.918 ± 0.103-Khong River, Thailand2008[11]
0.814 ± 0.089-Tuaran, Salut River area, Malaysia2008[11]
<0.029–0.195-Glatt River, SwitzerlandSeptember 2006[23]
1.94–32.853,540–32,430Wuhan urban lakes, ChinaOctober 2005[17]
<0.1–1.444–567Rieti district, Italy2002, 2003[12]
0–0.24-Danube RiverAugust–September 2007[24]
0–1.40-Tributaries of Danube RiverAugust–September 2007[24]
-<20–2,830Danube River2007[20]
<0.029–0.233-Ria de Aveiro, PortugalAugust 2006[25]
LOD-0.015LOD-1,750Great Lakes coastal wetland, Cootes Paradise, Canada2001, 2002[26]
75.2–179.65,460–119,100Lake Donghu, ChinaApril 2003[15]
0.036–33.231-Pearl River Delta, China2005, 2006[10]
0.152–13.757-Thessaloniki, Greece2005–2006[27]
0.015–0.386-Seine River Estuary, France2002–2004[9]
<LOD-0.770-Hessisches Ried region, Germany2003–2006[28]
<LOD-0.511-Donggang River, Taiwan2002[16]
0.1–0.575–340Cuyahoga River, Ohio, USA[29]
<0.210<350Minnesota lakes, USA2008[30]
0.068–0.326-Glatt River, Switzerland2004[31]
-107–16,198Pearl River system, China2006–2007[18]
-13–225Bay of Cadiz, Spain2002[32]
-31–21,885Pearl River Delta, China2006–2007[19]
3.1Dianchi Lake, China[33]
<LOD-1,364Upper Danube River2006[34]
47–192Venice lagoon, Italy2001–2002[35]
LOD: Detection limit. NA: not analyzed.
Table 2. NP levels in air.
Table 2. NP levels in air.
NP levels (ng/m3)LocationDetected timeReference

19.2 (1.5–69)6.1 (0.1–14)Hudson River Estuary, USA
10.2 (0.9–56)9.8 (0.3–51)Sandy Hook, USAJune–October, 1998[64]
2.5 (0.2–8.1)5.6 (1.8–23)Liberty Science Center, USA
6.9 (nd-56)5.4 (0.067–51)Sandy Hook, USA
2.6 (nd-17)3.8 (0.23–23)Liberty Science Center, USAJune–December 1998[66]
13 (0.13–81)0.55 (0.020–6.4)New Brunswick, USA
1.60–16.5-Urban site of Thessaloniki, GreeceJanuary–February 2007[63]
0.15–1.00.0017–0.117NE-Bavaria, GermanyMay–November 2001[65]
0.22 (0.055–0.42)0.040 (0.010–0.12)GKSS Research Centre, Germany-
0.056 (0.029–0.11)0.010 (0.005–0.017)North Sea-[67]
About 0.01–0.1-North SeaFebruary–March 2004[68]

Share and Cite

MDPI and ACS Style

Mao, Z.; Zheng, X.-F.; Zhang, Y.-Q.; Tao, X.-X.; Li, Y.; Wang, W. Occurrence and Biodegradation of Nonylphenol in the Environment. Int. J. Mol. Sci. 2012, 13, 491-505.

AMA Style

Mao Z, Zheng X-F, Zhang Y-Q, Tao X-X, Li Y, Wang W. Occurrence and Biodegradation of Nonylphenol in the Environment. International Journal of Molecular Sciences. 2012; 13(1):491-505.

Chicago/Turabian Style

Mao, Zhen, Xiao-Fei Zheng, Yan-Qiu Zhang, Xiu-Xiang Tao, Yan Li, and Wei Wang. 2012. "Occurrence and Biodegradation of Nonylphenol in the Environment" International Journal of Molecular Sciences 13, no. 1: 491-505.

Article Metrics

Back to TopTop