Nonylphenol in the Environment: A Critical Review A.J. Porter and N. J. Hayden

Department of Civil and Environmental Engineering University of Vermont Burlington, VT 05405

Abstract

Nonylphenol (NP), an anaerobic breakdown product of a widely used nonionic surfactant

nonylphenol ethoxylate (NPEO), is currently of environmental concern because of its toxicity,

estrogenic properties and widespread contamination. Nonylphenol has been detected in surface

and ground water, sediments, aquatic organisms, wastewater effluent, air, and human food.

Although concentrations of NP are typically low, the potential risk to human populations and

ecosystems at these concentrations is still unclear and is a topic of considerable debate. A better

understanding of the fate and transport of NP and its parent compound, NPEO, is needed for

determining the associated risk to the environment and human health posed by these compounds.

This paper presents a critical review of the literature and issues related to NP in the environment,

including the associated gaps, conflicts and research needs in this area. The similarities and

differences between NP and historical contaminants of interest such as PCBs and current

contaminants such as pharmaceuticals are noted, as is the need for analyzing the life cycle of

chemicals in the environment prior to widespread use. Finally, issues of risk and how these guide

policy as related to NP are discussed.

Key words: nonylphenol ethoxylate, estrogenic compounds, wastewater, nonionic surfactants

Introduction

Low concentration contamination from nonylphenolic compounds is nearly ubiquitous in

the environment. Nonylphenol (NP), which has numerous isomers, is of particular concern

because it is persistent, toxic to aquatic organisms, and a potential endocrine disruptor. Effects

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on human and ecosystem health due to low concentration exposure to NP are poorly understood

and open to considerable debate.

Widespread low concentration environmental contamination by persistent synthetic

organic compounds is not a new problem. Polychlorinated biphenyls (PCBs), DDT, and

polyaromatic hydrocarbons (PAHs) are well-documented historical examples (Neilson, 1994).

However, nonylphenols (NP) are different in that they typically are not released directly into the

environment, but rather are formed as the anaerobic biological breakdown products of widely

used nonionic surfactants, nonylphenol ethoxylates (NPEO), that are directly discharged.

Nonylphenol ethoxylates show neither the toxicity nor the estrogenic effects of NP.

Nonylphenol is another example that supports the notion of requiring a thorough life cycle

analysis of synthetic compounds and their fate in the environment before mass production and

widespread use.

The basic structure of nonylphenol is shown in Figure 1. The side chain has nine carbons

and can be attached to phenol at different points on the ring, thus producing different isomers.

Each is named according to the position of the chain attachment. For example, the NP shown in

Figure 1 is referred to as 4-NP. The carbon chain may be straight or branched in a variety of

configurations, and each different branching pattern of the carbon chain can also represent a

different NP isomer. Commercial production of NP for the making of NPEO results in a mixture

of NP isomers. One of the most common commercial forms of NP is 4-NP with a branched side

chain, and it is often used in experimentation and in the analysis of environmental samples

(Giger et al., 1984; Hellyer, 1991; Ekelund et al., 1993; Sweetman, 1994; Bokern et al., 1998;

and Sekela et al., 1999).

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Nonylphenols have been found in water samples from numerous locations worldwide

(Blackburn and Waldock, 1995; Ahel et al., 1996; Rudel et al., 1998; Snyder et al., 1999; Potter

et al., 1999; and Kuch and Ballschmiter, 2001). They have also been found in a variety of other

media including, sediments (Bennie et al., 1997; Bennett and Metcalfe, 1998; and Marcomini et

al., 1990), air (Dachs et al., 1999; and Van Ry et al., 2000), fish and mollusks (Lye et al., 1999;

Keith et al., 2001; Ferrara et al., 2001) and even human food (Guenther et al., 2002).

Quantifying low concentrations of NP in environmental samples can be difficult and

requires specialized handling and processing techniques (Lee, 1999). Specifically, care must be

taken to prevent loss of NP through volatilization or adsorption during collection and transport.

Different analytical processes, including various gas chromatography methods (Lee, 1999 and

USEPA, 2000) and high performance liquid chromatography methods (Lee, 1999), are required

to measure nonylphenolic compounds. At present, there are no standard methods for sampling,

storage and analysis of NP in environmental samples. This lack of standard protocols may result

in quality assurance/quality control (QA/QC) issues that need to be considered when reviewing

NP results.

The most common route of NP entry into the environment is through wastewater. The

nonionic surfactant group, NPEO, is typically used in domestic liquid laundry detergents,

industrial liquid soaps and cleaners, cosmetics, paints, and as the dispersing agents in pesticides

and herbicides (APE Research Council, 2001). Because of its extensive use in cleansers, most

NPEO are discharged to the sewer system and make their way into wastewater treatment plants.

Under anaerobic conditions such as those found in sewers, sediments, and certain treatment

operations at wastewater treatment plants, NPEO are biodegraded to NP. Understanding the

lifecycle of NPEO and NP in the environment is still an area for research.

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Nonylphenol and NPEO have been measured, typically in the microgram per liter range,

in the influent and effluent of many wastewater treatment plants (Brunner et al., 1988; Lee and

Peart, 1998; and DiCorcia et al., 2000). Generally, concentrations have been found to be higher

in the influent than in the effluent, indicating that some of these compounds are being “lost” at

the wastewater treatment plant (WWTP). However, the mechanisms for NP and NPEO loss are

not always clear. Nonylphenol is hydrophobic and it would be expected to preferentially adsorb

to biosolids. For example, Rudel et al. (1998) found concentrations of NP greater than 1000

µg/L in numerous septic tank samples collected from Cape Cod, Massachusetts.

Although there is general agreement that widespread, low concentration environmental

contamination by NP and NPEO exist, the level of risk to humans and the environment posed by

this contamination is currently under considerable debate by researchers, chemical

manufacturers, and regulators.

Two basic approaches exist for dealing with environmental contamination from the

discharge of chemicals such as NP into the environment. The first approach is to prove the safety

of a chemical and its potential byproducts prior to widespread use and discharge. The second

approach is to use the chemical unless the chemical’s toxicity and risk can be clearly proven.

European countries have taken the first approach and restricted or banned the use of NPEO since

the 1980s because of its potential environmental toxicity issues (Renner, 1997).

The United States has taken the second approach and allowed its use because toxicity and

problems at low concentrations have not been clearly proven. The problem with the second

approach is that the effects of low concentrations of contaminants in the environment can be so

complex and difficult to determine that clear scientific proof of toxicity may never absolutely be

determined, even though they cause environmental or human harm.

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This paper presents a critical review of the literature and issues related to NP in the

environment, including the gaps, conflicts and research needs that exist in this area. The

similarities and differences between NP and historical contaminants of interest such as PCBs and

current contaminants such as pharmaceuticals, are noted as is the need for analyzing the life

cycle of chemicals in the environment prior to widespread use. Finally, issues of risk and how

these guide policy as related to NP are discussed.

Critical Issues

The three critical issues making NP in the environment of particular concern are:

1. NP and its parent compound NPEO are nearly ubiquitous in the environment;

2. NP and NPEO life cycles indicate long term, continued, environmental contamination;

and

3. NP has been shown to be toxic to aquatic organisms and an endocrine disrupter in higher

animals, and therefore poses a risk to humans and the environment.

Ubiquitous distribution of NP in the environment. Concentrations of NP and its parent

compound NPEO have been measured worldwide in surface waters, sediments, sewage, the

atmosphere, aquatic organisms, and even in typical human food products.

Table 1 presents NP concentrations reported in the environment from a number of recent

studies. While the table is not exhaustive, it provides a cross-section of the range of samples and

concentrations reported in the literature. Generally, NP is present in very low concentrations in

aqueous solutions as shown in the surface water samples. Non-detect to the microgram per liter

range have been reported. This is expected based on NP’s low solubility and hydrophobicity

(discussed in detail in the next section). Biosolids and sediment samples often show higher

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concentrations than aqueous samples. In some of the solid samples, concentrations up to parts

per million were reported (Brunner et al. 1988).

It is important to note that when determining the impact of hydrophobic chemicals such

as NP in the environment, it is crucial that samples be taken in the media where the hydrophobic

chemicals would likely be found. For example, since NP adsorbs to solids, detecting low

concentrations of dissolved NP in a river without testing the associated sediment could

drastically underestimate the amount of total NP contamination.

Surveys of fish have reported concentrations of NP from non-detect to 180 ng/g (Lye et

al., 1999 and Keith et al., 2001). A survey of mollusks in Italy found NP at concentrations from

67 to 696 ng/g (Ferrara et al., 2001). These higher concentrations of NP in mollusks may be due

to the fact that the mollusks had more interaction with the sediment where NP is likely to be

associated than would the fish.

In addition to detection in the natural environment, concentrations of NP and its parent

compound NPEO, have also been measured in wastewater treatment plants (WWTPs). These

surveys have demonstrated the ubiquitous distribution of NP and NPEO in WWTPs, the

tendency of NP to be concentrated on solids, and the dependence on oxygen for the formation of

different NPEO end products.

Concentrations of NPEO in the influent and effluent have ranged from 29 to 415 µg/L

and non-detect to 332 µg/L, respectively (Brunner et al., 1988; Giger and Ahel, 1991; Lee and

Peart, 1995, 1998; Snyder et al., 1999; and DiCorcia et al., 2000). Concentrations of NP are

typically lower in the influent and effluent and have ranged from 0.8 to 22.69 µg/L and 0.171 to

37 µg/L, respectively (Brunner et al., 1988; Giger and Ahel, 1991; Lee and Peart, 1995, 1998;

Snyder et al., 1999; and DiCorcia et al., 2000). Concentrations of NP associated with biosolids

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have been measured in the hundreds of milligrams per kilogram range (Giger et al., 1984;

Sweetman, 1994; and Sekela, 1999).

Several studies confirm that WWTPs are the likely source of many NP in the

environment. Bennie et al. (1997) concluded that the concentrations of nonylphenolic

compounds in natural waters and sediments are highest near the outfalls from wastewater

treatment plants. Bennett and Metcalfe (1998) also determined that NP sediment concentrations

were high near urban and industrialized areas.

During wastewater treatment, NP can be biologically formed from NPEO when

anaerobic conditions are encountered. Some studies also suggest that NP can be biodegraded

under aerobic conditions (Ekelund et al., 1993 and Hesselsoe et al., 2001). Because wastewater

treatment can result in the production and elimination of NP, it is often difficult to determine the

NP degradation efficiency of WWTP based on the concentrations of NP and NPEO in the

influent and effluent. In addition, NP can volatilize and preferentially adsorb to solids making it

even more difficult to determine if biological treatment is even occurring. In spite of this

problem, influent and effluent concentrations are often used as an indicator of treatment efficacy.

Clearly, a better approach to quantifying NP fate in WWTPs is needed.

Nonylphenol has also been measured in atmospheric samples. Dachs et al. (1999) were

the first to report the occurrence of NP in the atmosphere, and found air concentrations ranging

from 2.2 to 70 ng/m3 in samples taken near the Hudson River. Another study from the same

laboratory, Van Ry et al. (2000), reports a similar range of NP atmospheric concentrations for the

same geographical region. These results suggest that volatilization should also be considered as a

mechanism when determining the environmental fate of NP; however, more work is clearly

needed in this area. Data for other geographical areas should be compiled since only one

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laboratory group has published atmospheric data. In addition to atmospheric sampling,

understanding the source of atmospheric NP, its form (i.e., whether it is in a free form or

adsorbed to particulates), and degradation mechanisms in the atmosphere such as

photodegradation is clearly warranted. Volatilization of NP especially during traditional aerated

wastewater treatment processes could result in long-range transport of NP through the air and is

an area for more study. Also the potential for human exposure to NP through the atmosphere is

largely unknown.

The difficulties associated with the sampling, handling and analysis of NP should be

noted at this point, especially since there is often limited information regarding these issues in

the literature. During sample collection, storage and processing, NP can be lost to volatilization

and adsorption leading to an underestimation of NP concentrations (Porter and Hayden, 2001).

Solid phase extraction is often used to concentrate NP from aqueous samples, and there are often

recovery issues with this technique. In contrast, carryover from one analytical sample to the next

could be a source for error in NP analysis (Porter and Hayden, 2001). Limits of detection for NP

vary from study to study, therefore samples that were non-detect for NP should be compared to

other studies carefully. Variability in the reported environmental concentrations of NP may be

partially attributable to the analytical methods used for quantification, but may also be a result of

the NP isomers present. Most researchers use 4-NP for analysis, and it is unclear what effect this

may have on quantifying NP mixtures. It is also difficult to obtain independent validation of

research laboratory results since most commercial U.S. laboratories either do not perform NP

analysis or have unacceptably high detection limits (Porter and Hayden, 2001).

In addition to widespread distribution of NP in the environment due to WWTP

discharges, there are other potential mechanisms for human exposure to NP. Guenther et al.

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(2002) found NP in samples of 39 typical human food items from a grocery store, as well as 21

baby foods, infant formulae and human breast milk. Guenther et al. (2002) measured NP

concentrations in all of the analyzed products ranging from 0.1-19.4 µg/Kg (based on fresh

weight).

The source of NP in the food was not clear. Additional research should be done to further

document NP concentrations in foods from other geographical locations, as well as to determine

how and when human food products become contaminated with NP. Due to the lack of a pattern

between the fat content, packaging type, and NP concentration, Guenther et al. (2002) concluded

that NP probably ends up in food products in a variety of ways and at different stages of food

handling and processing. Since NP is hydrophobic (log Kow = 4.48) it would be expected that

higher concentrations would be found in high fat foods, but no correlation was detected.

Guenther et al. (2002) found that there was no connection between the type of food packaging

and the NP content, in spite of the fact that foods packaged in tris(nonylphenol)phosphate

(TNPP) amended plastic could be expected to have higher NP concentrations due to leaching of

NP from plastic wraps. The two highest NP concentrations measured in the survey were in

apples (19.4 mg/Kg) and tomatoes (18.5 mg/Kg). Guenther et al. (2002) speculated that these

concentrations could be the result of exposure to pesticides that use NPEO as dispersing agents.

Ubiquitous contamination by NP is reminiscent of worldwide contamination by other

synthetic organics, including PCBs (polychlorinated byphenyls), MTBE (methyl tert-butyl

ether), and more recently pharmaceuticals, fragrances and caffeine. A recent study by Kolpin et

al. (2002) screened 139 U.S. streams in 1999-2000 for 95 contaminants including; NP,

pharmaceuticals, detergents, fragrances and other commonly used chemicals. They found many

of these compounds in 80% of the streams surveyed. Even in streams in the most pristine areas

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(those least likely to be influenced by agricultural, industrial or domestic wastes), caffeine and

triclosan (the active ingredient in anti-bacterial soaps) were detected. Some of the most

frequently detected compounds included, NP, steroids, insect repellents, and triclosan. Although

many substances were detected, most of the individual concentrations were very low (less than 1

µg/L).

All of these examples highlight the often uncontrollable and unpredictable behavior of

anthropogenic chemical releases into the environment. While the effects of these various

chemicals in the environment or on human health are often not clear, it is clear that human

releases are resulting in widespread contamination by hundreds, perhaps thousands, of different

long-lived chemicals. It is often only a matter of looking for them.

NP Life Cycle. In order to understand the lifecycle and fate of NP, it is important to consider

the properties and behavior of NP, as well as its relatively benign parent compound, NPEO. The

production of NPEO involves the reaction of NP with ethylene oxide to form a surfactant

composed of an ethoxylate chain (hydrophilic group) and an NP (hydrophobic group). During

production, a mixture of 4-NP isomers with branched hydrocarbon chains is typically used to

form the NPEO.

Biological degradation of NPEO results in a series of transformations that shorten the

ethoxylate chain. Figure 2 summarizes the aerobic and anaerobic biological degradation

pathways for NPEO. Under aerobic conditions, NPEO degrades to NPEO with shorter chained

ethoxylate groups or to NPEO with carboxylated ethoxylate and/or carbon chains. Concurrently,

the nine unit hydrocarbon chain can also be shortened to form other alkylphenolic substances.

Under anaerobic conditions, like those associated with sludge digestion in sewage plants, NP

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tends to be formed as the final product. In general, the shorter the ethoxylate chain, the more

hydrophobic, persistent, and toxic the substance becomes. Understanding the rates of conversion

in natural systems and WWTPs is an area where more research is currently needed.

Laboratory studies have demonstrated the possible degradation steps for NPEO under

aerobic environmental conditions. Historically it was thought that NPEO had the ethoxylate

chain shortened to one or two ethoxylate groups, and that the ethoxylate chains were

subsequently carboxylated (Manzano et al., 1999). Based on this understanding of the

mechanism of NPEO transformation, studies on NPEO and its breakdown products focused on

measuring the concentrations of short-chained NPEO and short-chained carboxylated NPEO in

environmental samples. Carboxylated NPEO with long ethoxylate chains were typically not

thought to occur and therefore not quantified in field samples. Jonkers et al. (2001), however,

showed that NPEO with long ethoxylate chains degrades first to NP with carboxylated

ethoxylate chains, forming long chained carboxylated NPEO and that the ethoxylate chains were

degraded next. Oxidation of the nonyl- chain was determined to occur at the same time as

carboxylation of the ethoxylate chain. By showing that NPEO with long ethoxylate chains may

first degrade to carboxylated NPEO with long chains, Jonkers et al. (2001) contend that previous

measures of NPEO breakdown products in natural waters may have been drastically

underestimated. In fact, although they found that greater than 99% of the NPEO was degraded

after 4 days, metabolites with carboxylated ethoxylate chains (carboxylated NPEO and other

carboxylated alkylphenol ethoxylates with less than nine carbons in the side chain) were still

present in the reactors 31 days after the experiment was started. This demonstrates that NPEO is

not as easily and ultimately biodegradable as once thought, although the initial NPEO is quickly

broken down.

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Maki et al. (1996) determined that NPEO with a carboxylated ethoxylate chain of length

one was the final breakdown product of NPEO under aerobic conditions in rivers. Similarly,

Maki et al. (1994) determined that under aerobic conditions the intermediate breakdown products

of NPEO were NPEO with a two unit ethoxylate chain and NPEO with a carboxylated two unit

ethoxylate chain. DiCorcia et al. (1998) found that the ethoxylate chain of NPEO was first

shortened and then carboxylated. The order of the breakdown or carboxylation of the ethoxylate

chain is open to debate. Under aerobic conditions, NP has not been found as an end product, and

many studies show that NP is not the primary degradation product of NPEO during aerobic

sewage treatment processes (Snyder et al., 1999; and Johnson and Sumpter, 2001).

Under anaerobic conditions, NP becomes a significant and important final product of

NPEO biodegradation. The ethoxylate chain of NPEO is sequentially shortened until NP is

formed as a final end product. Nonylphenol is very resistant to further anaerobic biodegradation

(Razo-Flores et al., 1996). For example, Ejlertsson et al. (1999) found that under anaerobic

conditions NPEO was degraded to shorter chained NPEO and then to NP. The NP was not

degraded further and is generally considered a persistent degradation product of NPEO.

Hesselsoe et al. (2001) also investigated the biodegradation of NP and found that NP added to

soil samples was not biodegraded after three months when the conditions were anaerobic.

However, laboratory studies show that NP can be biodegraded in aerobic situations. For

example, Ekelund et al. (1993) found that under aerobic conditions, after a 58-day trial,

microorganisms had mineralized 50% of the NP added to a seawater/sediment sample.

Additionally, Hesselsoe et al. (2001) found that the half-life of NP in soil under aerobic

conditions was 3 to 6 days.

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Understanding NPEO degradation in anaerobic environments is important because these

conditions are common during wastewater transport and treatment. Because of the extensive use

of NPEO in detergents, most NPEO eventually become part of the wastewater stream. This

represents a significant source of NP to the environment as the anaerobic breakdown product of

NPEO. Anaerobic conditions are purposely created during certain wastewater treatment

processes such as anaerobic sludge digestion and in anaerobic pretreatment tanks (septic tanks).

Additionally, areas of anaerobic conditions are common throughout the wastewater treatment

train, especially in settling basins, corners of tanks, and within flocculated particles. Anaerobic

zones may also occur in sewage lines, storm drains, and pipes as the wastewater is in transit to

the treatment facility. Thus, NP may be produced during wastewater treatment and transport,

since anaerobic areas are common. Field studies have supported this idea. Brunner et al. (1988)

did not find NP accumulation during aerobic sludge digestion, but significant NP accumulation

occurred during anaerobic sludge digestion.

Recent studies have also shown that even NPEO is not effectively removed or quickly

mineralized in traditional activated sludge wastewater treatment plants. Lee and Peart (1998)

found that there was a 53% elimination of alkylphenolic compounds with wastewater treatment

and that the majority of nonylphenolic compounds in the effluent were carboxylated NPEO.

Likewise, DiCorcia et al. (2000) concluded that under aerobic conditions, 66% of the influent

NPEO was converted to carboxylated NPEO. These results run contrary to the Alkylphenol

Ethoxylates Council’s claim that NPEO is very biodegradable (APE Research Council, 2001).

In fact, this claim is misleading, because although the initial NPEO is quickly transformed, the

intermediate degradation products are not easily mineralized, and it is possible that these

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intermediates can be converted to NP if the conditions become anaerobic (Maki et al., 1996;

Thiele et al., 1997; Potter et al., 1999; and Ejlertsson et al., 1999).

In addition to demonstrating that NP can be formed during anaerobic processes, the

WWTP studies found that NP in the waste stream preferentially adsorbs to the solids in the

system. Giger et al. (1984) examined anaerobically treated sewage biosolids and found a

significant accumulation of NP in the biosolids. Sekela (1999) found significant accumulation of

NP on the solids near WWTP discharge points indicating that either NP is formed during WWTP

treatment, or that NP enters WWTPs in the waste stream and is not effectively degraded during

treatment. Sweetman (1994) reported NP in anaerobically digested biosolids from two

wastewater treatment plants at mean concentrations of 638 and 326 mg/Kg.

The fate of NP in the environment is primarily determined by characteristics of NP such

as its resistance to biodegradation, hydrophobicity, and moderate volatility. Some of the basic

properties of NP are listed in Table 2. Variations in the values reported in the literature may be

due to evaluation of different NP isomers.

The high octanol-water partition coefficients (Kow) for NP are indicative of its

hydrophobic nature and this influences its environmental fate and transport. The log Kow of NP

is similar to that of anthracene (4.5) and other PAHs. Nonylphenol will tend to adsorb to

sediment in the environment and attach to solids during wastewater treatment. For example,

Ejlertsson et al. (1999) found that NP and short-chained NPEO adsorbed to the biosolids in

anaerobic experimental reactors. For the case of digested biosolids, they measured >50% NP

adsorbed to the solids at the end of the experiment. Giger and Ahel (1991) measured the

concentrations of NP in sewage plants and determined that 44% to 48% of the NP in the system

was adsorbed to the biosolids. They also experimentally determined the partition coefficient for

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NP to biosolids as 1800 L/Kg. Similarly, Ahel et al. (1994b) found the ratio of the concentration

of NP in river sediment to the concentration of NP in river water ranged from 364 to 5100. As

expected, sorption to solids was related to the organic matter content of the solid phase. The high

organic content of sewage solids also means a high adsorptive capacity. Experiments on the fate

of NP during sewage treatment should take into account the organic content of the solids in the

system in order to make valuable analyses.

Nonylphenol has a low solubility (Table 2), but because it contains phenol its solubility is

slightly dependent on pH. The hydroxyl group on phenol can dissociate at high pH resulting in a

slightly higher solubility as pH increases (Hellyer, 1991). Environmental factors such as pH,

temperature, total solids, aeration/turbulence, and surface area will influence NP dissolution.

Because NP is both lipophilic and resistant to biodegradation there is a potential for

accumulation in the tissues of higher animals. Predicted bioconcentration factors (BCF) and

bioaccumulation factors (BAF) are given in Table 2 for NP. The BAF is the ratio of the

concentration of a chemical in an organism to the concentration of that chemical in water at

equilibrium, and is based on uptake from the surrounding environment as well as from food. The

BCF is a similar ratio, but is only based on uptake from the surrounding environment and does

not include uptake from food. A BCF or BAF greater than 1000 is generally considered to

exhibit a risk due to bioaccumulation. A review by Servos (1999) rates the risk of NP to

bioaccumulate as low to moderate based on a survey of published data for BCF and BAF that

range from 0.9 to 3400. Some chemicals with Kow values similar to that of NP are known to be

biomagnified in the environment as is demonstrated by their very high BCF. For example, the

BCF of DDT is 34,000, 1,400 for anthracene, 3,200 for fluoranthene, and 6,100 for pyrene

(Savannah River Site, 2002). Biomagnification is a rare occurrence and does not seem to apply

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to NP, although NP does have many characteristics in common with known bioaccumulative

compounds.

Nonylphenol is moderately volatile which leads to concern of NP entering the

atmosphere from the aqueous phase. While its vapor pressure is relatively low, its Henry’s law

constant is in the moderately volatile range. Substances with Henry’s law constants between 10-5

and 10-7 atm-m3/mole are considered moderately volatile (Lyman et al., 1990). The Henry’s law

constant is the ratio of the concentration of a compound in air to the concentration of the

compound in water at equilibrium conditions. Henry’s law constants for NP are given in Table 2

and range from 1.55x10-5 to 4x10-5 atm-m3/mole.

Photodegradation may also play a role in the breakdown of nonylphenolic compounds in

the environment. Ahel et al. (1994a) found a half-life for aqueous NP at the water surface with

10-15 hours of light equivalent to midday, summer sun to be 1.5 times greater than in the water

20-25 cm below the water surface. It was also found that the rate of photolysis for NP increases

in the presence of organic matter. These results indicate that degradation of NP by UV light may

be a significant elimination method, but that other environmental factors and interferences need

to be considered in conjunction. It was unclear from the experimental methods used by Ahel et

al. (1994a) whether or not volatilization was accounted for in the photodegradation experiment.

If volatilization was not considered as a loss mechanism, the loss of NP by photodegradation

could be severely overstated. More research in this mechanism is clearly needed.

The importance of wastewater treatment in influencing the fate and transport of NP in the

environment should be noted. A recent review by Johnson and Sumpter (2001) notes that

activated sludge wastewater treatment for the removal of endocrine disrupting chemicals may not

be very effective. Not only is there the potential for NP formation at various stages in the

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process, there is also the potential for accumulation on the solids and volatilization that needs to

be better understood.

One of the most common conventional wastewater treatment systems for large

communities is the activated sludge system shown schematically in Figure 3. An activated

sludge treatment plant typically includes preliminary processes such as screens and grit

chambers, settling as primary treatment, and an aeration basin and secondary clarifier as

secondary treatment. Wastewater often enters the treatment plant from sewers in an anaerobic

state and preliminary processes are used to remove the gross solids from the waste stream. In

primary treatment, quiescent conditions are created in a clarifier in order to settle solids.

Secondary treatment is a biological treatment process whereby microorganisms metabolize the

soluble organics in an aerated tank. After aeration, a secondary clarifier is used to settle the

microorganisms. Depending on the waste characteristics, tertiary treatment may be implemented

to remove excess nutrients. The effluent is disinfected before discharge to the receiving water

source. In addition to the treatment of the wastewater itself, the byproduct biosolids from the

primary and secondary clarifiers require disposal. Sludge digestion, often anaerobic, is designed

to reduce biosolids volume. Dewatering may also be used to reduce the final biosolids volume

for disposal.

The potential for NP production is related to those locations and processes where

anaerobic conditions dominate including the sewer, clarifiers, and anaerobic digestion. Once

formed, NP has the potential to volatilize from the open tanks especially those that are aerated.

Lee et al. (1998) state that volatilization from the surface of primary and secondary systems in a

wastewater treatment plant are important fate mechanisms for compounds with moderate to high

volatilities. Nonylphenol falls within this range indicating that volatilization, even from standing

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waters needs to be considered as a significant pathway that determine the fate and transport of

NP both within wastewater treatment plants and the environment at large. Porter and Hayden

(2001) examined volatilization as a loss mechanism for NP during wastewater treatment and

concluded that up to 3% of the NP in wastewater influent could be released to the atmosphere.

As previously noted, NP will adsorb to biosolids and accumulate on the digested sewage

biosolids. Because of land application issues, investigations on the fate of NP in planted systems

(Bokern et al. 1998) and in the soil (Hesselsoe et al., 2001) have been forthcoming. Hesselsoe et

al. (2001) focused on the persistence of NP in biosolids amended soils and again demonstrated

that oxygen is the determining factor. They showed that it would take more than a year to reach

aerobic conditions within a two-centimeter diameter sludge aggregate and that until such

conditions were achieved, the NP associated with the aggregate would be persistent. Bokern et

al. (1998) determined that plants have the ability to uptake NP indicating that there is the

potential for NP accumulation in plants grown on biosolids amended soils and that

microorganisms are crucial to the mineralization of NP. The amount and rate of NP

incorporation into plant material was low (up to 4.2% of the NP added to sterilized soil was

detected in the shoots after 4 days) although this still demonstrates a potential risk. If NP were

to be incorporated into plants, NP could become part of the food chain which represents another

route of exposure for humans. Uptake through plants is one of the methods that could explain

the NP contamination that Guenther et al. (2002) found in human food and should be more

thoroughly evaluated.

Risk. One major area of debate is whether widespread low concentration environmental

contamination with NP presents a significant risk to humans or environmental health. The

determination of risk is based on a combination of toxicity and exposure data that takes various

18

priorities into account. Based on both exposure potential and NP toxicity, the degree of risk

posed by NP can be critically examined.

One factor contributing to NP risk determination is the exposure scenario and includes

such things as routes of exposure, duration of exposure and exposure concentrations. Clearly

understanding these issues are important for assessing human risk. For example, air routes of

exposure and drinking water concentrations are not clearly known. Also, while human

foodstuffs have been shown to contain NP (Guenther et al., 2002), it is still not clear what the

exposure concentrations are or where the food contamination originates. Possible mechanisms

for NP to enter the human food web include: through leaching from plastic packaging;

accumulation on crops as the NPEO found in many pesticides break down; and from

bioaccumulation due to land application of sewage biosolids or irrigation with sewage effluent.

On the other hand, there is documented environmental contamination showing definite routes of

exposure for aquatic life (Table 1).

Besides exposure scenarios, the toxicity of the compound is also an important factor in

determining risk. In a review by Servos (1999), acute toxicity of NP to fish, aquatic

invertebrates and algae was reported to occur in the range of concentrations from 17 to 3000

µg/L. Levels for no observable effect during chronic exposure to NP in fish and invertebrates

were reported at levels as low as 6 and 3.7 µg/L, respectively (Servos, 1999). The chronic and

acute toxicity levels for aquatic organisms were lower than the levels reported for some surface

waters in Europe(Table 1), and therefore, NP was restricted. Both toxicity and exposure

pathways were demonstrated. Renner (1997) notes that levels of NP in European waterways are

typically higher than those reported for the United States, although the cause of these differences

19

is not clearly known. They may be attributable to the significantly larger amount of dilution of

wastewater effluents in the United States as compared to those in Europe.

Nonylphenol has also been implicated as an endocrine disrupter in higher animals (Soto

et al., 1991; and Gimeno et al., 1997). Endocrine disrupters are chemical compounds that

interfere with the hormone system of an animal. Endocrine disrupters bind to a receptor site

instead of the naturally intended hormone. In some cases the endocrine disrupter mimics the

effect of a hormone, and in other cases the disrupter fills a receptor site thereby blocking the

intended hormone action. In either case, the effect is especially harmful during prenatal

development and differentiation.

There is significant evidence that NP acts as an estrogen mimic. Soto et al. (1991) found

that human breast cancer cells that require estrogen for proliferation would multiply in the

presence of NP. The estrogenic effects of NP were further confirmed in tests on live rats (Soto et

al., 1991). Gimeno et al. (1997) studied the effects of alkylphenols (AP), including NP, on the

sexual differentiation of male carp. Their results confirm that AP can function as hormone

disrupters at the whole animal level (Gimeno et al., 1997). Additionally, in a review by Servos

(1999), examples of NP acting as an estrogen mimic in fish are given for concentrations as low

as 10 µg/L.

The exact mechanism by which NP interferes with hormone activities is not clearly

defined, but one theory attributes the effect to the similar physical molecular structures of one

isomer of NP and an estradiol (Thiele et. al., 1997). As shown in Figure 4, certain patterns of

branched carbon chains on an NP molecule closely resemble the structural shape of estradiol

(estrogen). This emphasizes that certain isomers of NP may be estrogen mimics while others

may not. There is a lack of understanding on how important and potent each different isomer of

20

NP is as an estrogen mimic, and this represents a significant gap in current understanding of NP.

If further research shows that only certain NP isomers are endocrine disrupters, then screening of

environmental samples should focus on quantifying the implicated isomers in order to determine

the true degree of risk.

Although concentrations of NP contamination are generally considered to be low in the

environment, these concentrations may in fact be dangerous due to cumulative effects of low

concentration exposure in combination with other endocrine disrupters (Silva et al., 2002). Also,

because the current understanding of environmental endocrine disruptors in human development

is still incomplete, they should not be dismissed solely because of their low concentrations. Most

of the identified endocrine disrupters are also hydrophobic making the risk for bioaccumulation

in animal fat a significant threat (Staples et al., 1998).

Due to the demonstrated aquatic toxicity, the estrogenic effects, and the persistence of

NPEO breakdown products in the environment, the use of NPEO have been banned or

voluntarily restricted throughout Europe since 1986 (Renner, 1997). In contrast, there are no

laws regulating NPEO use or discharge in the United States. The lack of action on the part of

environmental regulators in the United States stems largely in part from the research conducted

by the Alkylphenol and Ethoxylate Research Council formed by the Chemical Manufacturers

Association to conduct studies on APEO (APE Research Council, 2001). To date this panel has

disputed all claims that NP concentrations in waterways of the United States are above

concentrations where a significant effect would be realized. The Alkylphenol and Ethoxylate

Research Council also contests the estrogenic potential of NP (APE Research Council, 2001).

There is clearly a need for continued research into the effects, even at low concentrations,

of synthetic organics such as NP in the environment. Recent research suggests that combinations

21

of low concentration endocrine disrupters may cause harmful effects (Silva et al., 2002). This

does not indicate synergy, but does advocate that the presence of mixtures of low concentration

contaminants can be harmful. Silva et al. (2002) found that when eight weakly estrogenic

compounds were combined, estrogenic effects were seen even though each of the component

compounds were present at concentrations that were individually too low to cause any

observable effects. This brings up the troubling issue of whether or not safety is ensured if all

individual chemical’s concentration is below the concentration that would cause an effect.

Determining the risk posed by endocrine disrupters should account for the additive effects of

these types of chemicals in the environment since the combination of many low concentration

contaminants could prove hazardous.

The risk assessment associated with synthetic organic compounds, like NP, is the central

factor in determining regulatory status. However, it is unclear where the burden of proof lies. In

North America, typically a compound must be determined to be toxic before regulations limiting

its use and release are instated. Currently there are no U.S. EPA guidelines or recommendations

on the safe concentrations of NP in waterways, although ambient water quality criteria for NP

have been under investigation since 1999 (USEPA, 1999). Conversely, in Europe, the use of

NPEO has been severely limited since 1986. The guiding principle for European regulators

seems to be opposite that of the U.S., thus requiring that compounds be proven safe. Erickson

(2002) explains that the U.S. and the European Union (E.U.) set different limits on

environmental assessment requirements. In the E.U., if the predicted concentration of a

compound in surface water exceeds 0.01 µg/L, then a risk assessment is required whereas in the

U.S. the concentration is 1 µg/L. This illustrates the more cautious approach of the E.U. to

environmental risk.

22

Alternate Strategies for Decreased Future Risks

There are several ways to decrease the risk associated with a particular compound. First,

the use and disposal of the compound could be restricted and regulated to decrease the quantity

that is discharged to the environment. In addition, new treatment methods could be developed to

remove the compound from the waste stream or the environment. These methods are often

supplemented by the introduction and use of a less toxic replacement for the compound. In the

case of NPEO, alcohol ethoxylates have been shown to be a less toxic substitute (Danish

Environmental Protection Agency, 2002; and Exxon Mobil, 2002); however, replacement of

NPEO may have an associated economic cost that users are unwilling to bear in light of the

unknown degree of risk represented by current usage. Developing effective, low-cost

alternatives to NPEO is warranted.

In the absence of NPEO substitution, new treatment strategies could be implemented to

remove NP from wastewater. Ultraviolet (UV) light could be a treatment option if NP is shown

to be photodegraded. Ultraviolet light is already a proven water and wastewater treatment

mechanism and could be developed to treat NP or other unwanted chemicals from wastewater.

Additionally, ozonation or filtration with granular activated carbon may have potential for

removing NP or other trace contaminants from wastewater since these processes have been

effective in eliminating pharmaceuticals from drinking water (Ternes et al., 2002). Economic

considerations would need to be ascertained.

Other technologies such as phytoremediation may also hold promise for treating low

concentrations of undesirable chemicals. Phytoremediation has been successful in treating some

organic contamination (Ferro et al., 1994; Schnoor et al., 1995; Hughes et al., 1997; Watanabe,

1997; and Schnabel et al., 1997). Bokern et al. (1996, 1997, and 1998) have studied the ability

23

of different plant species to uptake and transform NP, although it is not yet clear if plants are

able to metabolize and degrade NP, or if they transport and bind NP. Either mechanism would

be helpful for removing NP from wastewater. Additionally, Adler et al. (1994) studied the

possibility of using plant root surface exudates to break down phenolic compounds, reinforcing

the idea that naturally based plant and microorganism systems may be used to remove NP from

aqueous sources.

Conclusions

Nonylphenol is a persistent, virtually ubiquitous contaminant in the environment that can

be toxic to aquatic life, and has the potential to be an estrogen mimic. Although NP is present at

low concentrations, the risks associated with long term exposure to low concentrations of

endocrine disrupters are largely unknown.

As analytical methods improve, the detection and quantification of more organic

contaminants in the environment becomes possible. Moving toward a more thorough cataloging

of the pollutants present in our ecosystems elucidates the true lifecycle of the synthetic chemicals

introduced to the environment. Understanding the ultimate fate of the manufactured chemicals is

essential in order to avoid situations analogous to DDT or PCB contamination. Furthermore,

more research needs to be done to determine the potential human and environmental health risks

posed by exposure to combinations of synthetic organics in the environment.

Acknowledgements

Credits. The authors wish to thank the National Science Foundation POWRE program for

financial support.

24

Authors. A.J. Porter is a doctoral candidate and N.J. Hayden is an Associate Professor at the

University of Vermont in the Department of Civil and Environmental Engineering.

Correspondence should be addressed to: A. Porter, 213 Votey Building, Dept. of Civil and

Environmental Engineering, University of Vermont, Burlington, VT 05405; email:

aporter@emba.uvm.edu.

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29

Watanabe, M.E. (1997) Phytoremediation on the Brink of Commercialization. Environ. Sci. Technol. 31, 182A-186A.

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Table 1. Nonylphenol Concentrations in Environmental Media.

Source Location NP Concentration Reference Biosolids 11 U.S. WWTP 0.0054 to 0.887 g/Kg LaGuardia et al., 2001 Biosolids 5 Canadian WWTP 0.137 to 0.470 g/Kg Lee and Peart, 1995 Digested biosolids Swiss WWTP 0.45 to 2.53 g/Kg Giger et al., 1984 Digested biosolids English WWTP 0.326 to 0.638 g/Kg Sweetman, 1994 Digested biosolids Vermont WWTP 2.42 g/Kg Porter and Hayden, 2001 Digested biosolids 5 New York WWTP 1.130 to 1.840 g/Kg Pryor et al., 2002 Digested biosolids 29 Swiss WWTP 78000 µg/L Brunner et al., 1988 Raw biosolids 29 Swiss WWTP 2850 µg/L Brunner et al., 1988 River sediment St. Lawrence River 0.17 to 0.72 µg/g Bennie et al., 1997 Septic tank contents Cape Cod, MA 1000 to 1500 µg/L Rudel et al., 1998 Lake sediment Upper Great Lakes ND to 37 µg/g Bennett and Metcalfe, 1998 Surface water 139 U.S. streams 50.6% samples > ND;

max. of 40 µg/L Kolpin et al., 2002

Surface water St. Lawrence River 24% samples > ND; max. of 0.92 µg/L

Bennie et al., 1997

Surface water Glatt River, Switzerland

84% samples > ND; max. of 45 µg/L

Ahel et al., 1994b, 1996

Surface water England and Wales ND to 12 µg/L Blackburn and Waldock, 1995

Surface water Lake Mead, NV ND to 1.14 µg/L Snyder et al., 1999 Surface water Trenton Channel,

MI 0.269 to 1.190 µg/L Snyder et al., 1999

Surface water Tampa Bay, FL ND Potter et al., 1999 Ground water Switzerland 0.09 + 0.05 µg/L Ahel et al., 1996 Ground water Cape Cod, MA 25% of samples > ND;

max. of 1.5 µg/L Rudel et al., 1998

Drinking water Switzerland <0.1 to 7.2 µg/L Ahel et al., 1996 Drinking water Germany 2 to 15 ng/L Kuch and Ballschmiter,

2001 Air Hudson River

Valley 2.2 to 70 ng/m3 Dachs et al., 1999

Air Hudson River Valley

ND to 0.81 ng/m3 Van Ry et al., 2000

Human food Germany 0.1 to 19.4 µg/Kg Guenther et al., 2002 Fish U.K. rivers 5 to 180 ng/g Lye et al., 1999 Fish Michigan waterways 41% samples > ND;

max. of 21.9 ng/g Keith et al., 2001

Mollusks Italian waterways 67 to 696 ng/g Ferrara et al., 2001

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Table 2. General Chemical Properties of NP. Property Value Reference Formula C15H24O Shiu et al., 1994 Molecular Weight 220.36 Shiu et al., 1994 Specific gravity 0.950 Verschueren, 1983 Boiling point (oC) 315 Verschueren, 1983 Freezing point (oC) -10 Verschueren , 1983 Melting point (oC) 42 Shiu et al., 1994 Vapor pressure at 25 oC (Pa) 0.3 Muller and Schlatter, 1998 pKa (estimated) 10.28 Muller and Schlatter, 1998 Log Kow > 4.0 Thiele et al., 1997 Log Kow 3.80 to >4.75 Romano, 1991 Log Kow 4.48 Shiu et al., 1994 Log Kow 4.48 Ahel and Giger, 1993b Log Koc 4.4 Porter and Hayden, 2001 Log Koc 4.7 Sekela et al., 1999 Solubility (mg/L) 6 Muller and Schlatter, 1998 Solubility (mg/L) at pH = 5, 7, 9 4.6, 6.24, 11.9 Hellyer, 1991 Solubility (mg/L) 5.43 Ahel and Giger, 1993a Henry’s law constant (Pa-m3/mol) 3.04 to 4.05 Dachs et al., 1999 Henry’s law constant (Pa-m3/mol) 3.65 Van Ry et al., 2000 Henry’s law constant (Pa-m3/mol) 1.51 Shiu et al., 1994 BCF (bioconcentration factor), trout 24-98 Lewis and Lech, 1996 BAF (bioaccumulation factor), lab fish <1 to 1250 Staples et al., 1998 BAF, lab invertebrates 1 to 3400 Staples et al., 1998 BAF, field values 6 to 487 Staples et al., 1998 Toxicity to aquatic organisms (µg/L) 20-3000 Staples et al., 1998

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OH

C9H19

4-Nonylphenol (4-NP): Figure 1. Molecular Structure of 4-NP.

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Nonylphenol ethoxylate (NPEO); (n = 1 to 40; n = 9 is most typical in commercial mixtures).

O-[CH2CH2O]n-H C9H19

O-[CH2CH2O]n-H

C9H19

OH

C9H19

Nonylphenol (NP); recalcitrant compound.

Short-chained NPEO; (n is 1 or 2).

O-[CH2CH2]n-CH2COOH

C9H19

C9H19

O-[CH2CH2O]n-H

NPEO with carboxylated ethoxylate chain. Short-chained NPEO; (n is 1 or 2).

anaerobic

aerobic

anaerobic

anaerobic

aerobic

aerobic

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Figure 2. Biological Breakdown Pathways for Nonylphenol Ethoxylates in the Environment.

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screen/grit

clarifier (1o)

clarifier (2o) out sewer Cl2 aeration

return activated sludge

sludge digestion

sludge digestion

Figure 3. Schematic of Typical Activated Sludge Wastewater Treatment Plant.

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4-Nonylphenol (one isomer):

17β-Estradiol:

Me

OH

HO

HO

Figure 4. Comparison of the Chemical Structure of Nonylphenol and Estradiol (adapted from Thiele et al., 1997).

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