Linking Water Quality to Human Health and

Environment: The Fate of Micropollutants

Patricia Burkhardt-HOLM

AUGUST 2011

Serial No.IWP/WP/No.3/2011

Working Paper Series

www.lkyspp.nus.edu.sg/iwp

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"All substances are poisons; there is none which is not a poison. The right dose differentiates

a poison and a remedy."

This early observation concerning the toxicity of chemicals was made by Paracelsus (1493-

1541). It serves as a good starting point for the discussion on micropollutants. It reminds us

that our goal should not be to achieve zero concentration in our water bodies – a situation that

is neither feasible nor economically efficient - but to discuss and move towards acceptable

quality standard for our water resources.

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CONTENTS

Acronyms ............................................................................................................................... 4

Abstract .................................................................................................................................. 5

1. INTRODUCTION ............................................................................................................ 6

2. OBJECTIVE AND STRUCTURE .................................................................................. 6

3. DEFINITION OF TERMS ............................................................................................... 7

4. TYPICAL MICROPOLLUTANTS ................................................................................. 9

4.1 General ........................................................................................................................ 9

4.2 Frequently detected classes of micropollutants ........................................................ 11

5. FATE OF MICROPOLLUTANTS ................................................................................. 15

5.1 Entry paths into the aquatic environment ................................................................ 15

5.2 Principal distribution in the environmental compartments water and soil .............. 17

5.3 Degradation in the sewage treatment plant .............................................................. 19

5.4 Transport and transformation in natural water bodies ........................................... 20

5.5 Uptake and distribution within the organisms ......................................................... 21

5.6 Bioaccumulation, biotransformation and elimination processes in the organisms .. 24

5.7 Modes of toxic action ................................................................................................. 26

5.8 Impact of micropollutants in the biological hierarchy ............................................. 27

5.9 Effects in the population and species community ..................................................... 28

6. RISK ASSESSMENT ...................................................................................................... 29

6.1 Collecting information on the occurrence and concentration of micropollutants ... 29

6.2 Quantifying exposure of various organisms .............................................................. 31

6.3 Determining toxic effects ........................................................................................... 31

6.4 Extrapolating effects .................................................................................................. 35

6.5 Prioritizing micropollutants ...................................................................................... 36

6.6 Laws and Regulations ................................................................................................ 37

7. SELECTED EXAMPLES OF COMPOUNDS AND THEIR FATE IN THE

ENVIRONMENT ............................................................................................................. 38

7.1 Fate of ethinylestradiol (EE2) .................................................................................... 38

7.2 Fate of nonylphenol (NP) ........................................................................................... 44

8. POSSIBLE MEASURES ................................................................................................. 46

8.1 Source control ............................................................................................................ 46

8.2 Separation of effluent streams ................................................................................... 47

8.3 Sewage Treatment ...................................................................................................... 47

9. CONCLUSION AND OUTLOOK .................................................................................. 49

Glossary of Terms .................................................................................................................. 52

References ............................................................................................................................... 56

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Acronyms

CEC contaminant of emerging concern

DDE dichlorodiphenyldichloroethylene

DDT dichlorodiphenyltrichloroethane

E1 estrone

E2 estradiol

EDC endocrine disrupting compound

EDTA ethylenediaminetetra-acetic acid

EE2 17α- ethinylestradiol

EEF estrogen equivalent factor

EOC emerging organic contaminant

LOEC lowest-observed-effect concentration

MEC measured environmental concentration

MP micropollutant

NOEC no-observed-effect concentration

NP nonylphenol

NPEO nonylphenol ethoxylates

OECD Organisation for Economic Cooperation and Development

PAH polycyclic aromatic hydrocarbon

PBB polybrominated biphenyl

PBDE polybrominated diphenylether

PBT persistent, bioaccumulative, toxic chemicals

PCB polychlorinated biphenyl

PCDF polychlorinated dibenzofuran

PCP personal care product

PEC predicted environmental concentration

PFOA perfluorooctanoic acid

PFOS perfluorooctanesulfonic acid

PNEC predicted no effect concentration

POP persistent organic pollutant

PPCP pharmaceuticals and personal care products

QSAR quantitaive structure-activity relationship

STP sewage treatment plant

TBT tributyltin

VTG vitellogenin

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Abstract

Micropollutants are chemicals of great concern because of their ability to potentially cause

adverse effects in organisms at concentrations as low as a few ng/L. This discussion paper

aims at raising awareness and contributes to the understanding of the fate and the potency of

these contaminants.

Micropollutants are found almost everywhere on earth, including water bodies, soils and

human food. The main groups of micropollutants are pharmaceuticals, personal care products,

pesticides and industrial chemicals. Released by industry, households, or agriculture,

micropollutants enter the environment and spread throughout the water ecosystem. Although

micropollutants may be partly degraded in the sewage treatment plant or in the water bodies,

they are usually not completely removed. After being taken up by aquatic organisms or

humans via contaminated water or food, micropollutants are transported to different tissues

within the organism. Depending on the properties of the micropollutants and the biology of

the target species, they may bioaccumulate, metabolize or cause adverse effects. These effects

may translate into alterations on a higher biological level such as disruption of the hormone

system, followed by impacts on reproduction, etc. Eventually, effects may translate into other

compartments of the ecosystem. The methods and tests available for assessing the risk for the

environment and for humans have some shortcomings and consequently, models and safety

margins have to be applied.

Today, endocrine disrupting compounds are considered the micropollutants of greatest

concern for adverse effects and, accordingly, two endocrine disrupting compounds

(ethinylestradiol and nonylphenol) are examined. Their pathways and fate in the enviroment is

elucidated. Finally, it can be assumed that a risk for adverse effects due to exposure to

micropollutants exists for aquatic organisms but not for humans. Since risk can only partially

be estimated due to a lack of data, the precautionary principle should be applied in order to

avoid or reduce potential threats. The discussion paper ends with an outlook and addresses

open questions and suggestions for potential measures to avert the contamination of our water

supply and food chain by micropollutants.

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1. INTRODUCTION

In view of increasing water scarcity, sustainable management of water resources has become

imperative. Humans - as well as all living creatures - depend on water. Drinking water is the

most vital resource for life and, thus, water quality standards for drinking water and personal

hygiene are especially high. Nonetheless, water quality standards should also strive to

safeguard aquatic life from chemical and other water-borne pollutants. Water utilized for the

elimination of sewage and transport of pollutants is one of many important uses of the

resource, which simultaneously places stress on the resource. Water quality can also be

aggravated when a large amount of water is extracted from natural water bodies - for

example, for irrigation purposes - since pollutants then concentrate in the remaining water and

may reach undesirable concentrations.

As in Singapore, where reclaimed water is used to complement surface water and imported

water (Tan Yong Soon et al. 2009), water sources used for personal needs are often

replenished directly or indirectly by treated wastewater. Accordingly, concern over synthetic

chemicals that can resist wastewater treatment and that may contaminate water sources has

increased; it presents an issue that should be studied carefully. While hundreds of chemicals

were found not to cause an effect on aquatic organisms in concentrations < 1 µg/L, some

chemicals have been found to cause effects at concentrations at or below the detection limit.

Accordingly, the occurrence and frequency of these chemicals present in water supplies must

be monitored and their potential risk assessed carefully. At present, knowledge on many of

these contaminants, with respect to their effects on humans, animals, and their fate in the

environment, is largely deficient.

2. OBJECTIVE AND STRUCTURE

The objective of this discussion paper is to examine the cycle of micropollutants (MPs) in the

environment, while addressing their effects on humans and aquatic life. It gives a short

introduction to the classification of MPs and presents typical MPs, their origin, and entry

paths into the environment. Focusing on emerging MPs that occur frequently in surface water,

it also provides a comprehensive overview of the effect and fate of MPs in different stages.

While it provides quantitative data on consumption, sales, use, and concentrations of MPs in

different water bodies, the focus of this paper remains qualitative. General principles of

uptake in plants, animals, and humans, bioaccumulation and biotransformation, as well as

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elimination and transfer in the food chain are explained. The paper also delivers insight into

the current state of knowledge on general principles of exposure and effect.

While many aspects of MPs are still unravelling, this paper serves to examine and explain the

relevant underlying biological interrelationships. In this context, a short overview on the

assessment of potential risks to human and environmental health is presented. How toxic

effects can be assessed, either by means of toxicity tests or with the help of modelling, is

another aspect addressed in this paper. Criteria for a prioritization of MPs are also described

in short.

Two examples of well-investigated MPs - ethinylestradiol and nonylphenol - will be

presented. Their fate and the known or potential effects will be explained. These compounds

will be traced from the anthropogenic source - via release or excretion - to their entry into

water bodies and through to their subsequent uptake by plants, animals, and humans. Risk of

these compounds to human and ecosystem health is, thereby addressed in this discussion.

In the final section, the basics of the precautionary principle will be taken into account for the

proposition of measures. In the conclusion, recommendations on where to focus and how to

proceed with further research are made.

3. DEFINITION OF TERMS

A pollutant is defined as “a substance that occurs in the environment at least in part as a result

of man’s activities and has a deleterious effect on living organisms” (Newman and Unger

2003). MPs are trace compounds that occur in small amounts in the environment. It is

assumed that concentrations as low as a few ng/L can cause health effects in organisms

(Schwarzenbach et al. 2006, Murray et al. 2010). Apart from man-made MPs, minute

quantities of naturally-occurring toxicants, such as phycotoxins (e.g. nicotine), mycotoxins,

and pyrethrins, are present and can affect organisms.

In contrast, macropollutants are compounds occurring in greater amounts (µg/L-mg/L) and

include acids, salts, nutrients, and organic matter (Schwarzenbach et al. 2006). However, a

clear distinction between ‘macro’ and ‘micro’ cannot be made since classification also

depends on the actual site and situation.

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The terms ‘emerging contaminants’, ‘emerging environmental contaminants’, ‘emerging

organic contaminants’, or ‘contaminants of emerging concern (CECs)’ are also relevant

within the discourse on MPs. These terms specify chemicals that have only recently been

analysed or identified in the environment and which are believed to cause adverse effects on

ecosystems and humans. Nevertheless, they are still insufficiently regulated or entirely

unregulated (Murray et al. 2010). In the international context, endocrine disruptors and

pharmaceuticals were the first emerging micropollutants in the 1980ies and 1990ies, followed

by disinfectants, sunscreens/UV filters, perfluorinated compounds and brominated flame

retardants. At present, compounds like pyrethroid pesticides, phytoestrogens, benzotriazoles

and nanoparticles attract increased attention. It is important to note that new hazardous

compounds continue to emerge from development and production, but that rapid advances of

analytical technologies are also contributing to the increasing number of known contaminants.

Originally, gas chromatography was an important technique enabling the measurement of

minute concentrations of uncharged, non-polar compounds with some volatility, such as

dichlorodiphenyltrichloroethane (DDT) and polychlorinated byphenyls (PCBs) in the 1960s.

To measure polar (water soluble) compounds like those in pharmaceuticals and personal care

products (PPCPs) with this method, a transformation into more volatile derivatives (i.e.

compounds of similar chemical structure) was necessary. With the advent of the liquid

chromatography-tandem mass spectrometry, a new tool was made available for measuring

ng/L quantities of polar organic substances in all types of water bodies and solid matrices

(soil, sludge, sediments) without derivatization (Ternes et al. 2004, Snyder et al. 2008).

Nevertheless, their detection remains a great challenge since their chemical variety is

immense and their concentration level is usually low. Furthermore, chemical analysis is costly

and requires a significant level of expertise.

Due to legislative environmental measures, the emission of conventional priority pollutants

was reduced in most countries, thereby shifting the focus to remaining pollutants, which are

typically those found in lower concentrations (Martin and Voulvoulis 2009). Accordingly,

some authors distinguish between ‘conventional pollutants’ - such as persistent,

bioaccumulative, toxic chemicals (PBTs), persistent organic pollutants (POPs), and other

bioaccumulative chemicals of concern (BCCs) – and so far unrecognized or ‘emerging

pollutants’ (Ellis 2006).

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4. TYPICAL MICROPOLLUTANTS

4.1 General Typically, MPs are synthetic compounds. These are generated in huge numbers via human

activity. About 20 million synthetic chemicals are known; their numbers increase by 1 million

each year. The number of synthetic chemicals that are commercially available is estimated

between 50,000 and 100,000; this number increases by about 1,000 per year. Due to

byproducts of manufacturing (e.g. dioxins) and the breakdown of products and metabolites,

the total number of synthetic compounds in the environment is likely to be greater

(Worldwatch Institute 2011). The worldwide annual production of synthetic compounds

amounts to more than 300 million tons (Schwarzenbach et al. 2003). However, a

comprehensive list on the production numbers, use, emissions, toxicological properties, and

environmental effects of all of these compounds is lacking.

A great variety of synthetic chemicals are components of pharmaceuticals for humans as well

as animals, personal care products (e.g. bath and shower products, disinfectants and insect

repellents), and many other products common in daily use (Tab. 1). After their release ‘down

the drain’ and into the rivers, many of them continue to exist as MPs. Even MPs that are non-

persistent may cause prolonged negative effects due to their continuous introduction into the

environment (Ellis 2006). In addition, large quantities of pharmaceuticals and food additives

stemming from intensive animal husbandry, pesticides applied in agriculture, as well as

industrial chemicals and their byproducts threaten water resources (Tab. 1).

Finally, accidents are also a major contributor to the presence of MPs as well as

macropollutants in surface water. The explosion of the Deepwater Horizon platform in 2010

presents an example of the potential risks and effects of accidents. It is estimated that 500,000

tons of crude oil were released into the Gulf of Mexico as a result of this accident

(http://www.zeit.de/news-nt/2010/7/16/iptc-bdt-20100716-428-25600522xml). Burning of the

crude oil led to contamination with polycyclic aromatic hydrocarbons (PAHs) in the water.

Furthermore, the surfactants used to mitigate the effects of the oil spill were actually classified

as toxic. Recent events - namely, the earthquake and the tsunami in Japan - present another

example of the risk and potential effects of accidents. Within the first six days of the accident,

concurrent with a large amount of radioactive water, a cocktail of toxic chemicals of unknown

composition and amount were released into the environment.

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Tab. 1 Important micropollutants, subgroups, examples and main adverse effect (EDC: endocrine disrupting compound)

Group Subgroup Examples Main adverse effect

pharmaceuticals(human)

synthetic hormones (contraception, hormone replacement therapy)

17α-ethinylestradiol (EE2)

EDC

painkillers ibuprofen EDC

diclofenac EDC

naproxen EDC

lipid regulators bezafibrate EDC

fenofibric acid EDC

clofibric acid EDC

ß-blockers propanolol EDC

antibiotics trimethroprim EDC, bacterial resistance

ciprofloxacin EDC, bacterial resistance

sulfamethoxazole EDC, bacterial resistance

antiepileptics carbamazepine EDC

psychostimulants caffeine moderately toxic

medical products

radiocontrast agents iopromide EDC

pharmaceuticals(veterinary)

growth hormones (anabolic steroids)

trenbolone acetate EDC

antibiotics sulfonamides, tetracyclines EDC

antibiotics chloramphenicol EDC

animal feed additives

growth enhancement arsenic roxarsone toxic, carcinogenic

personal care products (PCPs)

sun / UV screens methylbenzylidene camphor (4-MBC) EDC, largely unknown

antioxidants and preservatives parabens (hydroxy benzoic acid) EDC

antimicrobials in cosmetics, household chemicals

triclosan EDC, persistent degradation products

nitro- and polycyclic musks tonalide (AHTN) EDC

disinfection by-products N-nitrosodimethyl-amine (NDMA) carcinogenic

insect repellants N,N-diethyl-metatoluamide (DEET) EDC

biocides herbicides alachlor toxic, carcinogenic

atrazine toxic

bentazone EDC 2,4 D EDC diuron EDC

insecticides carbaryl toxic

dieldrin EDC

dichlorodiphenyltrichloroethane (DDT)

EDC

fungizides vinclozolin EDC biocides tributyl tin (TBT) and compounds EDC

industrial chemicals

flame retardants, plastics brominated diphenyl ethers (PBDE) EDC, carcinogenic

surfactants and metabolites nonyl phenol (NP) EDC

bisphenol A EDC

fluoro-surfactants perfluorooctanesulfonic acid (PFOS) EDC, carcinogenic

perfluorooctanoic acid (PFOA) EDC, carcinogenic

anti-oxidants (food additives) butylated hydroxyanisole (BHA) carcinogenic

plasticizers diethyl phthalate (DEP) EDC

coordination complex ethylene-diaminetetraacetic acid (EDTA)

toxic, mutagenic

corrosion inhibitors benzotriazole (BTSA) toxic

coolants PCBs EDC, carcinogenic, mutagenic

incineration products dioxins, PAHs EDC, carcinogenic, mutagenic

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4.2 Frequently detected classes of micropollutants

Pharmaceuticals are substances used in medication including, but not limited to,

contraceptives, painkillers, lipid regulators, beta-blockers, antibiotics, tranquilizers,

psychostimulants, and impotence drugs. The United States, Japan, and Germany are the

world's largest markets for pharmaceuticals (http://www.buyusainfo.net/docs/x_3931631.pdf).

Over 200,000 different drugs and health products are registered in the US, Canada, European

Union, Japan, and other countries (www.drugs-about.com). In the EU, about 3,000 different

pharmaceutical substances are available for human use (Ternes et al. 2004). The number of

pharmaceuticals available is steadily increasing; nearly 900 medicines and vaccines are

currently being tested in the battle against cancer (www.phrma.org). In 2010, 54 new

pharmaceutical products were authorized in the EU (www.ema.europa.eu); in 2009, 25 new

pharmaceuticals were approved in the US (www.fda.gov). From 2004 to 2009, worldwide

sales of pharmaceuticals increased by 6.7% per year to more than $800 billion US$ (North

America: 5%; Europe: 7%; Asia, Africa, and Australia: 14%; Japan: 4%; Latin America:

11%). The global pharmaceutical market is estimated to grow by 5% to 7% in 2011 (IMS

2010). Veterinary pharmaceuticals, like antibiotics and growth hormones, are widely used in

animal husbandry and aquaculture. The Electronic Animal Drug Product Listing Directory of

the US Food and Drug Administration includes more than 1,430 approved products.

In wastewater and surface water, human and veterinary pharmaceuticals typically occur in

low range concentrations of ng/L (e.g. synthetic hormones), low to high range concentrations

of ng/L (e.g. ß-blockers, antibiotics and antiepileptics), and up to concentrations of hundreds

of ng/L (e.g. painkillers, lipid regulators) (Tab. 2). Thus far, about 200 different

pharmaceutical substances have been detected in surface water (Molander et al. 2009).

In a study of 20 drinking water utilities in the US, finished drinking water in at least half of

the utilities contained the antiepileptics carbamazepine and dilantin, and the antianxiety drug

meprobamate in the pg to ng range (Snyder 2008). Occurrence of pharmaceuticals in samples

of finished drinking water have been reported in Europe (e.g. Adler et al. 2001), but scientific

studies over a longer time scale and a larger geographical range are not available. In India,

pharmaceuticals were found, for example, in wells in the vicinity of pharmaceutical

production sites (Fick et al. 2009). In contrast, micropollutants in bottled mineral water stem

from leaching of compounds from the plastic bottles (Wagner and Oehlmann 2009).

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Personal care products (PCPs) include hair care, skin care, decorative cosmetics, oral hygiene,

perfume and fragrance, bath and shower, deodorant, soap, sun screen, insect repellent, and

disinfectant products. More than 11 billion individual personal care products are sold in the

US annually (http://www.cosmeticsinfo.org/fdapartner.php). In 2009, the global market for

personal care products approached $300 billion at the retail level. All cosmetic products may

contain more than 1,500 different ingredients (www.cosmeticsinfo.org), while fragrances may

contain 3,163 ingredients (http://www.ifraorg.org/en-us/Ingredients_1). Problematic for the

environment are, for example, certain active ingredients such as tensids, polycyclic musks,

UV-screens and antimicrobials. Often, the same active ingredients are, among others,

components of household cleaning products (e.g. detergents, laundry aids, dish soap, furniture

care, floor polish, carpet cleaners, and car cleaners). Concentration measurements of PCPs in

water bodies are in the ng/L up to low µg/L range (Tab. 2).

Two of the most frequently detected MPs in surface waters belong to the pharmaceuticals and

personal care products - together also called PPCPs: (1) ibuprofen - a non-steroidal, anti-

inflammatory drug that is also the third most popular drug in the world both with respect to

sales and use - and (2) triclosan - an antimicrobial disinfectant found in a wide array of

personal care products, household cleaning products, and plastics (Carr et al. 2011).

In the agricultural context, over two million tons of pesticides per year are used worldwide

(US EPA 2011a); they comprise 900 active ingredients in 40,000 commercial preparations.

The Environmental Protection Agency (EPA) estimates that the use of pesticides doubled

between 1960 and 1980 (http://www.answers.com/topic/pesticide). These pesticides enter

waterways mainly by diffuse discharge, however, concentration peaks in the water courses

due to cleaning of application equipment on farms have been reported as well. Pesticides in

surface waters usually show concentrations < 1 µg/L (Tab. 1). Common pesticides are the

herbicides alachlor and atrazine, the fungicide vinclozolin, the insecticide DDT, and the

biocide TBT. Most countries (including India, China, Malaysia etc.) have legislation banning

certain pesticides; other pesticides are restricted, for example DDT may only be applied for

malaria control in India. Nonetheless, banned chemicals are often used if there are

unsufficient controls. In addition, problems arise when permitted pesticides are used without

adhereing to application and/or disposal instructions (Dewi and Pertiwi 2006).

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Tab. 2 Micropollutant classes and examples of concentrations in wastewater and surface water (STP: sewage treatment plant; M: median; na: not available; concentrations of a specific micropollutant in wastewater and surface water mostly do not stem from the same study or the same area; sources: Adler et al 2001; Clara et al. 2011; Murray et al. 2010; Pal et al. 2010; Porter and Hayden 2002; Scharf et al. 2002; Tejon et al. 2010; Yu et al. 2009; citations therein)

Group Subgroup Examples of

substances Examples of concentrations reported in wastewater (range, median or mean) (ng/L)

Examples of concentrations reported in surface water (range, median or mean) (ng/L)

pharmaceuticals(human)

synthetic hormones 17α-ethinyl-estradiol (EE2)

7 0-5

painkillers ibuprofen 24 - 314 0-360

diclofenac 370 1-150

naproxen 810 4

lipid regulators bezafibrate 1810 16-363

fenofibric acid 81 45

clofibric acid 130 3-248

ß-blockers propanolol na 0-20

antibiotics trimethroprim 150 0-212

ciprofloxacin 111-370 23-1300

sulfamethoxazole 332 15-2000

antiepileptics carbamazepine 250 3-157

PCPs antimicrobials triclosan 4.8 140

nitro- and polycyclic musks in fragrances

tonalide (AHTN) 410-1800 up to 1200

insect repellant DEET na 60

pesticides herbicides atrazine 7 2000

insecticides dieldrin na 180

industrial chemicals

flame retardants, plastics

brominated diphenyl ethers (PBDE)

1.9 0.01

surfactants nonyl phenol (NP) 800-22690 800

fluoro- surfactants

perfluorooctanesulfonic acid (PFOS)

8-375 6

anti-oxidants (food additives)

BHA na 50

plasticizers diethyl phthalate (DEP)

na 200

corrosion inhibitors benzotriazole (BTSA)

na 230

A number of MPs are industrial chemicals, for example, flame retardants (brominated diphenyl ethers), surfactants (nonylphenol), fluorosurfactants in coatings (PFOS), anti-oxidants (butylated hydroxyanisole), plasticizers (phthalates), corrosion inhibitors (benzotriazole), or incineration products (dioxin). Although PCBs were banned in many countries in the 1970s and 1980s as well as by the Stockholm convention in 2004 (http://chm.pops.int/Convention/tabid/54/language/en-US/Default.aspx), they are still found almost everywhere in the environment. Typical concentrations of industrial MPs in surface water are below 10 µg/L (Tab. 2). Heavy metals are also occasionally categorized as MPs, although these may occur in very high amounts at sites, such as mining areas. Arsenicals, for example, are used as feed

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additives, stimulating growth and improving bird pigmentation (Arikan et al. 2008). They occur in surface waters in the low µg/L range (Tab. 2). Information on usage and occurrence of recently emerging MPs in aquatic systems such as nanoparticles, biosimilar medications (e.g. human insulin), or nutraceuticals (e.g. huperzine - a bioactive food supplement) is lacking to a large extent. Although emissions of some toxicants are declining due to regulations, bans, and substitutes, a growing world population creates an increasing demand for personal care products, industrial products, and food from intensive agriculture. Furthermore, there is a legacy of more than 200,000 tons of abandoned pesticides that may end up in the environment. This issue is of particular concern in Africa (http://www.un.org/ecosocdev/geninfo/afrec/vol15no1/ 151envir.htm). About one third of these abandoned pesticides are thought to be persistant organic pollutants.

5. FATE OF MICROPOLLUTANTS 5.1 Entry paths into the aquatic environment

The principal point source of most MPs is the effluent of sewage treatment plants (STPs) (Fig.

1). Further discharge entering the environment directly or mainly via STPs, also stem from

manufacturing, processing and distribution, hospitals (Ternes et al. 2004, Ellis 2006), and

aquaculture.

Fig. 1 Entry paths of micropollutants into the environment Sewer exfiltration was studied showing that leakages from aging urban drainage systems

result in peaks of the pharmaceuticals diclofenac, clofibric acid, iopamidol, and ibuprofen in

groundwater at depths of 25cm to 50cm with concentrations of up to 10% of that in the sewer

(Ellis 2006).

Prescribed drugs are usually only partly metabolised in the human body; remaining fractions

including metabolites are excreted via urine and faeces and reach the STPs via domestic

sewer systems. ‘Down-the-drain’ disposal of pharmaceuticals, when left-over or expired

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drugs are flushed down the toilet, is another problematic pathway of entry of the active

substances.

When drugs are disposed to dumping sites, leakages may contribute to groundwater

contamination. Dumping sites are regarded as ‘ecological time bombs’ (Walker et al. 2006),

because they contain many unknown pollutants in which case the fate of the pollutant is

unpredictable with regard to leakage from corroding containers, chemical reactions, or

degradation.

Non-point (diffuse) emission of biocides from materials, such as roofs, paints from buildings

also adds to the contamination by MPs. Here, factors like UV irradiation and contact with

water enhance the leaching (Schoknecht et al. 2009). As well, surface run-off from

agricultural areas treated with pesticides (e.g. pest control, vector control), takes place and is

an important additional entry route. After heavy rainfall, a variety of pesticides can be

regularly found in drains and waterways. In addition, natural organic compounds with toxic

properties, such as estrogenic mycotoxins (Bucheli et al. 2008), may enter the water bodies

via runoff from neighbouring fields. A fraction of these non-point pollutants may infiltrate the

soil and reach the groundwater system (Murray et al. 2010).

MPs from veterinary pharmaceuticals and food additives are released into the environment

from animal excretions in urine and faeces; in addition, washoff by rain of topical treatments

of livestock raised on pastures carries them into the environment (Boxall et al. 2004). When

solid or liquid manure is used as fertilizer and applied to the fields, veterinary pharmaceuticals

and food additives contaminate the soil and may be released via runoff into the water bodies.

Application of sewage sludge to agricultural fields is an inexpensive and easy solution to

handle the increasing quantities of sludge, hereby improving soil fertility and reducing the

amount of synthetic fertilizers used. However, persistent organic pollutants and other MPs

characterized by considerable bioaccumulation potential are present in varying concentrations

in the sludge and may lead to contamination of soil. In Spain, a study investigating the human

health risk due to long-term sewage sludge application on agricultural soils showed that

concentration levels of contaminants in food (vegetables, meat, and milk) were linearly

correlated to concentrations of the contaminants in the soil (Passuello et al. 2010). Due to the

risk of spreading contaminants and pathogens, the use of sludge in agriculture has been

banned in several countries (e.g. Switzerland) but is still allowed elsewhere. In Singapore,

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sludge produced in secondary waste water treatment is used as a soil additive for reclaimed

land (http://www.un.org/esa/agenda21/natlinfo/countr/singapor/natur.htm).

Soil as well as sediments can store high amounts of contaminants. Especially under anaerobic

conditions (no or low oxygen content), degradation may be poor. In cases of flooding,

contaminated soil and sediments may be dislocated and may cause pollution in former pristine

areas. Furthermore, leachates and run-off from contaminated soil may reach surface water or

ground water.

Aerial deposition occurs when volatile and gaseous contaminants or contaminants bound to

airborne particles are washed out by rain or snowfall. Pesticides are often applied by aerial

spraying, giving rise to volatile and gaseous contamination, especially under warm

temperatures. The problem of spray drift arises when pesticides in a gaseous state are

transported over great distances by air and then precipitate elsewhere. However, when

associated with particles or droplets, they are most likely only to move over short distances

before falling down. In order to minimize air pollution during pesticide spraying operations,

environmental factors such as wind speed and direction and humidity have to be taken into

consideration. Apart from pesticides, other small volatile contaminants, such as halogenated

hydrocarbons are also transported via air, either due to diffusion with concentration gradient,

or with thermal diffusion. Other airborne MPs are generated by combustion of fuels (PAH,

PCB, dioxin), either from factories and incineration plants, private burning, or from

combustion engines of motor vehicles, ships, and airplanes.

5.2 Principal distribution in the environmental compartments water and soil

The environmental fate of chemicals is basically determined by two factors: (i) the properties

of the compound itself, and (ii) the conditions of the surrounding system.

Based on the chemicals’ intrinsic physicochemical properties and reactivity levels, they can

be characterized as polar chemicals, ionisable organic chemicals, or chemicals with a high

number of functional groups (Schwarzenbach et al. 2006). Depending on these properties,

chemicals either go into solution, are in a suspended form as droplets, or adsorb to particles

(Fig 2). For any substance in water, its octanol-water coefficient (KOW value) is used for

predicting its partitioning between particles and water. Octanol and water are immiscible.

Chemicals that are commonly found in the octanol phase (lipophilic chemicals) are more

likely to adsorb to particles in natural water bodies than those which are found more in the

water phase after they have been added to a standardized octanol-water mixture. Thus,

lipophilic organic MPs of high KOW (e.g. PBDEs, nonylphenol) are strongly bound to soil,

sediment, or sludge, whereas polar MPs (hydrophilic compounds of low KOW, e.g. the

herbicide atrazine) tend to dissolve in water. Sorption further depends on salinity, pH-level,

and other factors. Sorption to suspended or settled particles, dispersion, and volatilization

causes a dislocation of the compound (Fig. 2).

In the soil or water, MPs may disappear due to volatilization or may be abiotically broken

down by hydrolysis, oxidation, isomerisation, and photochemical breakdown. Breakdown

usually results in a loss of toxicity, but sometimes the products are equally or even more toxic

as in the case of the isomerisation of malathion to isomalathion. The half-life of the MP

depends on the soil and water properties, the type of compound, and the temperature.

Besides the basic properties, the dynamics in the system are responsible for a high variability

in the concentration of MPs in water bodies. Seasonal variation of compounds that are applied

or used differently during the course of a year – such as antibiotics, pesticides, or application

of sewage sludge or organic fertilizer – result in fluctuating concentrations.

Fig. 2 Fate of micropollutants in water

19

5.3 Degradation in the sewage treatment plant

In sewage treatment plants (STPs), MPs are either dissolved in water or adsorb to sludge (see

5.2) and are then broken down by chemical or biochemical processes, for example by

hydrolysis, oxidation, or photodegradation. In biological wastewater treatment, many MPs are

readily biotransformed enzymatically. Accordingly, a general decrease in concentrations of

compounds from STP influent to effluent can be seen.

Most ingredients from personal care products are highly lipophilic and significantly sorb onto

sludge and sediments. Understanding the relevant processes in STPs for removing PPCP was

the aim of the EU project POSEIDON (Ternes et al. 2004). Pharmaceuticals of high

molecular weight and high Kow, such as analgesics (e.g. acetoaminophen, ibuprofen,

mefenamic acid) and estrogens (Tab. 3), can be removed between 76% to 96 % by sorption to

activated sludge, i.e. sludge containing large quantities of microorganisms (Yamamoto et al.

2007). Removal rates for polycyclic musk fragrances (PCMs) range from 72% to 86% and

92% to > 99% for UV screens (Kupper et al. 2006). In another study, overall removal rates

(sorption plus biodegradation) between 30% and 75 % were reported for all PPCPs tested

(except iopromide). In more detail, the removal rate was 65% for 17b-estradiol, between 40%

and 70% for anti-inflammatories, around 60% for the antibiotic sulfamethoxazole, between

30–50% for musks, and between 70% and 90% for fragrances (Carballa et al. 2004). Large

differences in removal rates between studies elucidate the importance of an optimal set-up of

the STP. Commonly, residues from STPs are then discharged to receiving water bodies where

a further transformation or a natural transport takes place.

20

Tab. 3 Removal of selected micropollutants in sewage treatment plants (sources: Beausse 2004, Fahlenkamp et al. 2006, Kameda et al. 2007, Pal et al. 2010, Snyder et al. 2003, Suarez et al. 2008, Syracuse SRC Interactive PhysProp Database 2011, Yu et al 2006, and citations therein; na: not available)

Group Examples of substances log Kow

natural biodegradation

STP removal (%) (adsoption and degradation, highest value reported)

pharmaceuticals 17α-ethinyl-estradiol (EE2)

4.2 persistent 98

ibuprofen 3.97 slow 95

diclofenac 4.51 rapid 75

bezafibrate 4.25 <20% 83

propanolol na slow 96

ciprofloxacin 0.4 >90% 83

carbamazepine 2.45 poor 45

medical product iopromide 2.05 na 85

animal feed additives

arsenic roxarsone -0.05 yes; metabolite more toxic

na

PCPs triclosan 4.76 yes 69

tonalide (AHTN) 5.7 not inherently biodegradable

90

nonylphenol 5.92 yes 95

diethyl phthalate 2.42 yes 95

EDTA -3.86 persistent na

benzotriazole (BTSA) 1.44 poor na

5.4 Transport and transformation in natural water bodies The degradation that occurs in natural water bodies again depends on various aspects, such as

aerobic and anaerobic conditions, temperature, etc., and may involve hydrolysis, oxidation,

photolytic degradation, or biodegradation (see, for example, nonylphenol, Chapter 7.2, Fig.

8). Principally, natural attenuation – or ‘self-cleaning capacity’ - is higher in streams than in

standing water bodies. This is due to the mostly higher oxygen content (i.e. aerobic

conditions) and turbulences leading to better mixing, and, as a consequence, a higher

degradation in flowing water bodies. However, there is a knowledge gap with regard to

degradation reactions and degradation products in surface waters (Helbling et al. 2010).

Degradation may take hours to weeks, again, depending on both the compound and the

physico-chemical conditions.

Besides degradation, dilution takes place in the natural water bodies. As consequence to

dilution, concentrations are lower the further they are removed from the discharge; thus, hot

spots are closer to the source of the discharge. A biological gradient can be generated with

21

sensitive species missing near the source of pollution and reappearing further downstream.

Sometimes, a high abundance of opportunistic species can be found in the vicinity of the

discharge.

Low, insufficient, or missing degradability also means that these compounds can be

transported via air or water hundreds or thousand of kilometres away and can harm the new

environment in which they are deposited. Whereas stable compounds in solution can cover

greatest travel distances in fast flowing rivers, particulate matter tends to fall to the bottom

and droplets may rise to the surface. If MPs are persistent, they can reach high concentrations

in the water bodies (e.g. the x-ray contrast medium iopromide, heavy metals). Also,

depending on the substance and physicochemical conditions, groundwater studies show that

PPCPs can survive very long time spans. For example, carbamazepine and primidone stay

intact for travel times of 8 to 10 years (Ellis 2006).

In oceans, transport over long distances is accomplished via currents, again depending on

properties of the chemicals as well as of the water. Large circular patterns of currents (gyres)

transport debris and pollutants all over the Atlantic and Pacific and may transport pollutants

from one continent to the next. The plastic gyres show how these movements work and give

an idea of what this might signify for humans and the environment. Indeed, plastics not only

present a risk when mistakenly taken up by marine organisms, they also act as material onto

which hydrophobic pollutants like PCBs, DDT, or PAHs may absorb (Endo et al. 2005).

5.5 Uptake and distribution within the organisms

Many microbes, plants, and higher organisms take up MPs (Tab. 4) and a transfer to various

internal sites takes place. If a MP molecule meets an organism, several prerequisites are

necessary in order for an effect to be generated: First, the chemical needs to be bioavailable,

which means it must be available in a form that can be absorbed and/or be taken up by the

organism. Little investigation has been done in order to determine to which extent pollutants

are taken up when they are bound to particles such as sediments or soil (Walker et al. 2006).

In general, uptake is dependant on properties of the molecules as well as on the organism’s

integument (surface) structure and physiology.

To be taken up by any organism, MPs must enter the cell across biological membranes and

other tissues. Passive diffusion across the membrane requires lipophilicity, since the

22

membrane itself is made up of a lipid layer. In other cases, carriers facilitate transport or

transport is mediated by receptors (specific, active transport). While most industrial chemicals

are non-ionized, the passage of charged ions such as amines, carboxylic acids, phenols, and

some pesticides through cell membranes is mediated by proteins called ‘proton pumps’.

(i) uptake by microbes

Microorganisms (also called microbes, which comprise all uni-cellular organisms, bacteria,

fungi, archaea, and protists, plankton, micro-algae etc.) are able to transform, degrade, or

accumulate a wide variety of compounds including hydrocarbons (e.g. PAHs), PCBs,

pharmaceutical substances, and metals.

(ii) uptake by and transport in plants

Higher aquatic plants are exposed to MPs via contaminated surface waters, terrestrial plants

by contaminated soil, reclaimed irrigation water, the application of pesticides, and

atmospheric deposition. Further, the application of organic fertilizers such as manure or

poultry litter (containing veterinary pharmaceuticals or arsenic food additives), or sewage

sludge (containing, for example, brominated diphenyl ethers) are further pathways of

exposure. Plants influence the fate of MPs by adsorption (Shi et al. 2010) and uptake.

Furthermore, they enhance microbial degradation of MPs by releasing co-metabolizing

compounds. In general, plant roots are the main site for uptake of chemicals from soil, but the

leaves – especially in aquatic species - may constitute another pathway. The uptake involves

equilibration between the concentration of the chemical in the aqueous phase within and

outside the plant root, followed by chemical sorption to lipids in membranes and cell walls

within the root. Lipophilic organic chemicals such as PAHs, chlorobenzenes, and PCBs are

usually more ready to partition into plant root lipids than hydrophilic chemicals (Briggs et al.

1983). Water and solutes are transported from the root into other plant compartments through

the vascular system by mass flow resulting from a pressure gradient created during

transpiration. Uptake efficiency depends on the compound, the plant species, and the root

lipids. MPs are transported to and stored in differing fractions in various tissues and organs,

including seeds and fruits. For example, the amount of arsen stored by rice was highest in

roots, followed by stem, leaf, husk, and then grain (Liu et al. 2009).

23

Tab. 4 Examples of uptake and reduction of micropollutants by plants Compound Plant species Uptake characteristics References pharmaceuticals (carbamazepine, triclosan)

soybean uptake + translocation Wu et al. 2010

pharmaceuticals (carbamazepine, sulfomethoxazole)

cabbage uptake + translocation (root > leaf/stem)

Herklotz et al. 2010

hormones (EE2, E2, E1)

algae, duckweed ca. 5% adsorbed, uptake unknown

Shi et al. 2010

animal food additive (arsenic roxarsone)

rice uptake + translocation (root >stem >leaf >husk >grain)

Liu et al. 2009

industrial chemical (nonylphenol)

lupin etc. uptake ca. 1.5% Bokern et al. 1998

combustion product (PAHs)

celery, rye grass, white clover

uptake < 2% + plant-promoted biodegradation

Meng et al. 2011

PBDE tobacco, nightshade

uptake + translocation (above-ground > shoots, fruits)

Vrkoslavova et al. 2010

(iii) uptake by and transport in animals and humans In principal, all animals are exposed to MPs; uptake was shown for various species ranging

from earthworms, which are a major food source for many higher organisms (Halford 2008),

to humans. Uptake depends on the size of the molecule and its polarity, as well as on the

structure and physiology of the exposed species. As in all organisms, uptake is either the

consequence of passive diffusion or active processes across biological membranes (Tab. 5). In

fish and amphibians, organisms that are charcterized by a permeable skin, many pollutants are

taken up via the epidermis (the outer layer of the skin) and the permeable gills. In mammals

and birds, however, the skin is reinforced with keratin, making it impermeable to most

chemicals and therefore, the main pathway is via ingested food and water.

In vertebrates and humans, transport via the blood often occurs by transport proteins. In egg-

laying species, a transfer of lipophilic compounds into the eggs is described. In mammals and

humans, a transfer of pollutants across the placenta into the developing embryo is known.

Lipophilic compounds – such as pharmaceuticals, PCBs, dioxins, polychlorinated

dibenzofurans (PCDFs), polybrominated biphenyls (PBBs), polybrominated diphenylethers

(PBDEs), DDT, and heavy metals – are secreted into the milk and, thereby passed to the

infants or offspring (Stefanidou et al. 2009). Micropollutants absorbed from the gut are

mainly transported to the liver.

24

Tab. 5 Uptake routes of micropollutants by organisms (adapted from Walker et al. 2006) Type of organism Route of uptake Sources of pollutant (aquatic) plants roots soil or ambient water

leaves droplets or particles, vapors

aquatic amphibians alimentary tract food, small amounts from water

skin water

fish skin water

gills water

alimentary tract food aquatic mammals and birds alimentary tract food, small amounts from water

terrestrial invertebrates alimentary tract food and water

cuticle (insects) surfaces

body wall (slugs, worms) soil

tracheae droplets and particles in air, vapour

terrestrial vertebrates alimentary tract food and ingested water

skin surfaces lungs droplets and particles in air, vapour

After being taken up by and distributed within organisms, MPs may interact with a molecule

or, more generally, a site of action. The internal concentration, for instance at the target site,

determines whether or not and to which extent an effect will be generated (see Escher and

Hermens 2004).

Various interactions between an exogenous molecule and the exposed organism are possible.

These interactions might lead to a transport of the exogenous molecule to the target organ or

tissue, where an effect might occur, or the compounds may accumulate or they may be

transformed and eliminated.

5.6 Bioaccumulation, biotransformation, and elimination processes in the organisms

The type and amount of a substance present in the organism determines if and to which extent

effects may occur. Low doses of chemicals may cause effects after a long exposure time or

after bioaccumulation. Bioaccumulation is the general process of storing compounds in

organisms. It is expressed as a factor resulting from the concentration in the organism relative

to the concentration in the environment. If the result of uptake is due to diffusion, in the case

of aquatic species from the surrounding medium, it is called bioconcentration. The extent as

well as the path of bioaccumulation depends strongly on the substance, the biochemistry of

the respective organism, its lipid content, and further parameters like the bioavailability of a

compound. For example, methylmercury has a much higher bioavailability than the inorganic

mercury. Most EDCs are known to bioaccumulate, but not all of them biomagnify (Ahel et al.

25

1993). Biomagnification is when accumulation of compounds occurs along the food chain, i.e.

increases with the ecological hierarchy. PCBs are an example for compounds that biomagnify.

Biotransformation summarizes all processes in which the original compounds are transformed

into compounds that can be eliminated by the organism. This depends again on the

compound’s properties and - in principle - leads to a loss of toxicity. In vertebrates, two

phases of biotransformation can be distinguished, which eventually lead to the synthesis of

more polar compounds that can be excreted more easily. For some xenobiotics, however, the

products of the first transformation step are more toxic than the parent compounds, such as

organophosphorous insecticides and carcinogens. The oxidation products of the latter are

highly reactive metabolites that bind to the DNA and lead to damages. However, their half-

life is so short that they are difficult to detect.

The capability to metabolize MPs is very different among the various organisms and taxa;

however, knowledge about this metabolism in different organisms is still in its infancy.

Algae are known to biotransform MPs, for example estrogens, certain phenols, naphthalene,

and diaryl ethers (Della Greca et al. 2008). Although more detailed studies are necessary,

metabolism of MPs in plants occurs mainly via enzymatic pathways. Various studies showed

that plants are able to reduce MPs via phytoremeditation (Herklotz et al. 2010, Shi et al. 2010,

Liu et al. 2009, Meng et al. 2011, Vrkoslavova et al. 2010, Bokern et al. 1998). However,

remediation is rather due to adsorption and microbial enhancement than to uptake (Tab. 4).

In animals, site, process, and velocity of excretion depends as well on the size and

lipophilicity of the chemical. The smaller and the more water soluble (polar) the compounds

are, the more easily they can be excreted. In fish, small molecules are excreted via skin and

gills. In amphibians, skin is the excretion route. Aquatic birds, reptiles, and mammals cannot

excrete via the skin as this organ is relatively impermeable to pollutants. As all vertebrates,

they excrete polar compounds and conjugates of high molecular weight (> 300 MW) via bile

and compounds of small molecular weight (< 300 MW) via urine (Walker et al. 2006) (Fig.

3).

Fig. 3 Pathway of micropollutants in fish 5.7 Modes of toxic action

In principle, three different modes of toxic action can be distinguished (Schwarzenbach et al.

2006). First, a disturbance of biological membranes, their structure or their function, is a

common mechanism. Since all organisms are characterized by biological membranes, this

non-specific mode of action is also called ‘baseline toxicity or narcotic toxicity’. When the

membrane function is disrupted, the organism shows a decreased capability to react to stimuli

and may even die. About 60% of industrial chemicals entering the aquatic environment are

believed to cause narcotic toxicity (Van Wezel and Opperhuizen 1995). Narcotic toxicity

needs relatively high concentrations, which are largely independent of the type of molecule

and the biological species.

‘Receptor-mediated toxicity’ is found when a chemical binds to a specific receptor. This

represents the largest group of MPs. Such contaminants belong to the endocrine disrupting

compounds (EDCs) interfering with the endocrine system. Such MPs may mimic natural

hormones and bind to the corresponding receptor instead. Two examples include

ethinylestradiol (EE2) and trenbolone. EE2 mimics the natural estradiol and binds to the

estrogen receptor; trenbolone mimics the natural testosterone and binds to the androgen

receptor. Other pharmaceuticals belonging to this group are the pharmaceuticals tamoxifen

27

(used in breast cancer treatment) and flutamide (used in prostate cancer treatment), the

pesticides atrazine, dieldrin and toxaphene, and the surfactant alkylphenol-ethoxalate.

‘Reactive toxicity’ is typical for electrophilic chemicals (e.g. acrylamide, ethyl acrylate,

acrolein, nitroaniline) interacting unselectively with nucleic acids and proteins found in

biomolecules and will not be discussed here (see Freidig et al. 1999).

Some MPs are listed as mutagenic (e.g. benzene) or genotoxic (e.g. the industrial or

household detergent EDTA, which was shown to be genotoxic in laboratory rats as well as the

plasticizer DEHP/diethylhexylphthalat). Furthermore, benzene, nitroso-compounds, butylated

hydroxyanisole, arsenic roxarsone, and some pesticides (e.g. alachlor) are carcinogenic.

5.8 Impact of micropollutants in the biological hierarchy The different modes of action might result in biochemical, physiological, morphological

changes (such as impaired cell metabolism) and may affect the next higher level in the

biological organisation, namely the function of organs and, finally the organism (Fig. 4).

While molecular effects on the first level are suspected to be most sensitive and provide the

earliest indications, their ecological relevance is typically low, since many molecular effects

are abolished by repair and protective mechanisms of the organism.

On the higher levels of biological hierarchy, there may be, among others, histological lesions,

effects on behaviour, reduced immune functions associated with increased susceptibility to

pathogens, adverse effects on reproduction, genetic makeup, or allergies. PCBs were found to

reduce fecundioty and offspring survival in many different species, including predatory birds

and marine mammals.

Fig. 4 Levels of biological hierarchy 5.9 Effects in the population and species community

When functions, such as growth, reproduction, or viability of individuals are affected, effects

on the population level may be manifested as well. They only emerge with a temporal delay,

depending on the time needed until reproduction takes place of the respective species under

investigation. Further, effects on the population level are mostly unspecific, though,

retrospectively unravelling the cause-effect relationship on this level of the biological

organisation is rarely possible.

Furthermore, due to the complexity of the food web (Fig. 5) - which involves smaller species

being eaten by larger ones - effects of micropollutants on certain species result in

unpredictable effects on the ecosystem. Accordingly, humans face a higher risk of exposure

due to consumption of seafood or fish than due to consumption of water.– as the most critical

compounds bioconcentrate along the food chain and reach higher concentrations in higher

levels of the hierarchy than in the water. In order to prevent a risk for species, ecosystems and

humans, levels of micropollutants in water have to be monitored.

When input of MPs ceases, for instance as a result of a ban on specific MPs, some systems

may recover without further action. For example, the concentration (body burden) of DDT

and PCBs in fishotters in the UK decreased after the ban of these chemicals. Also, fish

populations may rebound from former exposure to EDCs, as in Karen Kidd’s experiments

with experimentally contaminated lakes in Canada, where the fathead minnow population

recovered again three years after stopping the exposure (Halford 2008).

Fig. 5 The aquatic food web

6. RISK ASSESSMENT Risk assessment includes (1) collecting information on the occurrence and concentrations of

toxicants in the environment, (2) quantifying exposure of various organisms (3) determining

toxic effects, and (4) extrapolating effects to other individuals, life stages, the population,

other species, etc.

6.1 Collecting information on the occurrence and concentration of micropollutants

(i) by measurement

Information of occurrence and concentrations of MPs in water bodies can be obtained from

measurements. However, with regard to the large number of MPs and the existence of

innumerable water bodies on earth, it is impossible to measure more than a minute fraction.

30

(ii) by estimation

In the absence of real-world data, concentrations can be predicted from a combination of

production, consumption or sale studies, data on elimination in STPs, and their physico-

chemical properties.

Production, consumption and sales studies provide preliminary information on the amounts of

MPs used. In a second step of such an assessment, the fractions of these released into the

environment (e.g. via excretion of pharmaceuticals, emissions of industrial and personal care

products, or pesticide run-off) is estimated.

Release into the environment after use may differ strongly. For example, the estimated release

of the pain-killer diclofenac into the environment is 16% (i.e. the amount excreted), while the

release of the corrosion inhibitor benzotriazole from dishwashing powder is 100 %.

Comparisons of consumption studies and concentration measurements may provide additional

information with regard to release estimations. In a Swiss study, it was found that the

percentage of biocides released to the environment in urban areas was higher than in

agricultural areas. Furthermore, the study revealed that the use of biocides in urban areas is

not restricted to any particular season – as is the case in agriculture (Wittmer et al. 2011).

It is also important to note that the formation of metabolites must be taken into consideration,

since these are often found in higher amounts than the parent compounds (Boxall et al. 2004).

In the case of diclofenac, 65% of the compound is excreted as metabolites. Although the

potency is only 20% of the parent compound, metabolites of other compounds may have have

higher potencies than the parent compound.

Predicting the fate of MPs also requires knowledge of how the elimination of compounds

occurs in STPs, which may vary strongly (see 5.3 Degradation in the sewage treatment

plant). Based on an extensive review of recent literature, the US EPA offers a downloadable

database that allows users to predict the fate of MPs in STPs, (US EPA 2010a). The result of

these different estimation steps (production, release, transformation, and elimination) is the

predicted environmental concentration (PEC).

31

6.2 Quantifying exposure of various organisms To quantify exposure of aquatic organisms, the concentration of the MP in natural water

bodies is typically determined. Due to physicochemical properties, however, not all of the

MPs in the water are taken up and are bioavailable. For this reason, it is much more advisable

to determine the body residue if possible. For human exposure, average daily water

consumption or food intake is used as a base.

6.3 Determining toxic effects (i) with living organisms

To determine toxic effects and the respective concentrations, traditionally, acute or long-term

field studies or laboratory tests with living organisms are used. Toxicity tests are carried out

with fish such as fathead minnows, rainbow trout and zebrafish, invertebrates such as water

fleas, algae, and bacteria, which are standardized by OECD, ISO, or country-specific

guidelines. With fish, flow-through tests and exposure times of 24 to 96 hours are commonly

performed to determine the effect concentrations on different endpoints such as the acute

lethal concentration at which half of the population dies (LC50), or physiological,

reproductive, and behavioural endpoints such as VTG induction, fertilization, sex ratio,

spermatogenesis, thyroid parameters, or histology (Tab. 6). Simpler and less expensive tests

with invertebrates use immobilization or reproduction as endpoints.

In addition to animal welfare considerations, there are several other disadvantages to

conducting toxicity tests with living organisms. For practical reasons, toxicity tests usually

use only a few different concentrations of the toxicant. As a consequence, reactions to other

exposure concentrations have to be extrapolated. This is especially problematic at very low

doses, where chemicals might sometimes cause unexpected effects. This is the case in

inverted U-shaped dose-responses, that is, lower doses induce a more profound effect than

higher doses (Markey et al. 2003). In contrast, hormesis is a concept describing the fact that

extremely low doses of a compound may prove to be beneficial to an organism, while higher

doses may have adverse effects. Also, traditional toxicity tests in aquaria only consider the

exposure pathway of compounds dissolved in water but in the real world, fish may

additionally ingest food contaminated with the same (or other) toxicant.

32

Toxicity tests do usually not distinguish between the effects of the synthetic chemical and its

metabolites or transformation products, which result from abiotic breakdown (e.g. photolysis)

or biodegradation by micro-organisms or by the plants, animals and humans by which they

were taken up. EU regulations (directive 91/414/EEC) require evaluation of all major

degradates formed at more than 10% of the concentration of the parent compound. Degradates

are all products of degradation, whether biotically or abiotically derived. Thus far, studies

performed have suggested a low risk of degradates to organisms in standard toxicity tests (e.g.

Daphnia, rainbow trout, earthworms) with exception of a few industrial substances such as the

surfactant nonylphenol (NP) as well as some pesticides showing sublethal and longer-term

effects (Osano et al. 2002).

33

Tab. 6 Examples of toxicity tests, possible species and endpoints Species common name Example of endpoint

fish

rainbow trout (Oncorhynchus mykiss), zebrafish (Brachydanio rerio), fathead minnow (Pimephales promelas), medaka (Oryzias latipes)

growth, death, deformities, reproduction, egg and offspring development, VTG induction

amphibians

leopard frog (Rana pipiens) VTG induction

molluscs

neogastropods imposex

invertebrates

water flea (Daphnia spp) immobilization, reproduction

algae Selenastrum capricornutum growth

bacteria Photobacterium phosphoreum luminescens

cell bioassays

E-screen breast cancer cells cell proliferation

YES-screen yeast species colorimetric and luminescent response

non-cellular assays enzyme-linked immunosorbent assay (ELISA) colorimetric response

enzyme-linked receptor assay (ELRA) luminescent, colorimetric response

(ii) with bacteria, cell cultures or bioassays

An alternative to toxicity tests with living vertebrates or invertebrates are bacterial tests, cell

cultures, and non-cellular bioassays. Bacterial bioassays are easier, cheaper, and faster than

those with invertebrates or vertebrates. Cell bioassays use living cells from animals, plants, or

humans to assess the toxicity of MPs. They aim to detect highly specific effects that are not

likely to be influenced by conditions of the whole organism.

Although they are excellent screening tools, comparability of different commercially-

available assays is sometimes low. In addition, extrapolation of results to the whole organism,

for the purpose of risk assessment, is still in its infancy. This is partially due to the fact that

cell culture systems usually lack the whole equipment of pathways necessary for the

biological efficacy. For example, enzymes required to convert the precursors of hormones to

the active hormone are not always present. This problem is especially important in studies on

the endocrine system, because this system is particularly dependent on feedback mechanisms

of the whole organism.

(iii) with models

34

Due to the aforementioned shortcomings, models are frequently used. Toxicokinetic models

may help in predicting the fate of organic chemicals in organisms. Several processes, such as

uptake, distribution, metabolism, and elimination, can be looked at independently. For

example, quantitative structure–activity relationships (QSARs) can be used to predict the

toxicity of MPs based on the physicochemical properties (Boxall et al 2004, see also above).

For toxicological assessment, the predicted no effect concentration (PNEC) is determined.

Usually, a PNEC value is estimated by dividing the lowest no-observed-effect concentration

(NOEC) for the most sensitive species by a safety factor (Pal et al. 2010). For risk assessment

of pharmaceuticals, for example, the EMEA guideline (EMEA 2006) “enforces” the use of

chronic toxicity data and requires long-term NOEC (e.g. three NOEC values from three

different trophic levels such as algae, Daphnia, and fish, applying an assessment factor of 10

to the lowest value). When a measured environmental concentration (MEC) (or a predicted

environmental concentration (PEC), based on estimates) and a PNEC are available, a risk

quotient is usually calculated dividing the measured environmental concentration (or PEC) by

the PNEC. Due to a lack of ecotoxicological data (e.g. NOEC values), however, this is not

always possible (Besse et al. 2008); thus, NOECs have to be replaced by values such as

lowest observable effect concentrations (LOECs) or toxicity thresholds (Lin et al. 2008).

When PECs surpass the PNEC - i.e. the ratio of predicted environmental concentration

divided by the PNEC is equal to or higher than 1, measures are needed to protect humans and

the environment from harmful effects.

The values used for PEC reflect average situations, while the reality might be quite different:

MPs may occur continuously in low concentrations (e.g. traffic, PAHs, residues from

pharmaceuticals, persistent degradation products of surfactants,), in low, but highly

fluctuating concentrations (pesticides, run-off from fields), and in high and highly fluctuating

concentrations (pesticides, biocides in first-flush events).

Furthermore, MPs often occur in the environment not as single compounds, but in mixtures

with many other chemicals. Whereas the individual component might be present in

concentrations too low to cause effects, additive or even synergistic effects can cause

detrimental impacts on organisms. This was demonstrated in several studies that showed the

cumulative impact of a mixture of chemicals had adverse effects on fish, while the

concentrations of the single compounds were too low to cause measurable effects (Brian et al.

35

2005). Accordingly, a mixture of 13 common pharmaceuticals disrupted the growth of

embryonic liver cells at very low concentrations (Pomati et al. 2006). Additive or synergistic

effects were reported for fish (Schnell et al. 2009), daphnids (Schmidt et al. 2005), amphipods

(Engraff et al. 2011), and black fly larvae (Overmyer et al. 2003). Such cumulative effects

cannot only be seen with chemicals sharing the same mode of action, but also with those

acting differently. It is currently discussed whether minute stress of a multitude of chemicals

are the real big environmental stress, possibly aggravated by other unfavourable

environmental conditions, such as temperature outside the preferred temperature range, noise,

competition, and others.

6.4 Extrapolating effects

Extrapolating effects from toxicity tests with certain species or models to effects at higher

ecological levels (population, other species, food chain) is probably the most difficult task.

A population of the same species comprises different life-stages of which some might be

more sensitive than others. Indeed, fetal exposure to certain pharmaceuticals is a cause of

concern. For example, agonists of the aryl hydrocarbon receptor (such as dioxins and PCBs,

mimicking the molecule naturally binding at this receptor) are very toxic to early life stages of

fish, whereas adult fish are 10 times less sensitive (Cook et al. 1993). The standard endpoints

do not provide sound information on the effects on the total population. If, for example, an

LC50 has been established, environmental concentrations at this level leading to the death of

half of the population could cause a total breakdown after a couple of generations. However,

opposite effects might also occur and mechanisms such as increased survival of remaining

progeny could compensate for such a breakdown.

Extrapolating effects from toxicity tests or models from one species to the next is also

difficult. Currently, 1.6 million different species are known; many more are not yet described.

Their sensitivity towards chemicals is different depending on their habitat, food, position in

the food chain, as well as their physiology (e.g. their biochemical properties in detoxifying

xenobiotics). More specific detoxification mechanisms are developed in vertebrates and

organisms with a longer lifespan than in invertebrates, meaning that more target sites are

available for disruption by MPs. Dioxins and PCBs, for example, are not very toxic to

invertebrates, which lack the aryl hydrocarbon receptor to which those compounds typically

36

bind (US EPA 2008a). Regarding pharmaceuticals, water fleas were the most susceptible

followed by fish and algae (Sanderson et al. 2004).

In addition, the fate of pollutants in the food chain is not well understood and can hardly be

predicted. A study by Hook and Fisher (1998) showed that zooplankton was affected when

feeding on phytoplankton that was exposed to silver concentrations as low as 0.05 µg/L, while

they were unaffected when exposed to the same concentration in water themselves. Many

studies found PCB burdens increasing from lower to higher levels in the food chain (e.g. from

Blue mussle to chink shell, butterfish and black guillemot) (Skarphedinsdottir et al. 2010).

6.5 Prioritizing micropollutants Since it is impossible to monitor all MPs that might occur, priority chemicals are usually

determined. Priority chemicals are those which are frequently detected and show a reasonable

potential to adversely affect aquatic life (US EPA 2008a). To protect aquatic life in the US,

for example, two criteria, based on maximum concentrations and on continuous

concentrations are defined to limit exposure. Accordingly, data have to be produced (e.g.

measurements of prioritized compounds have to be carried out at ‘hot spots’) for example

from STPs and drinking water wells.

According to a very comprehensive, qualitative risk assessment, antibiotics, sex hormones,

cardiovascuolar compounds, and anti-neoplastics were ranked in descending order with

respect to their potential severity for human and environmental health hazard. Antibacterial

resistance is assessed as the most significant human health hazard, whereas the largest non-

target organism hazard appears to be endocrine disrupting compounds in wildlife (Sanderson

et al. 2004).

Whereas aquatic organisms are exposed their entire lives via skin, gills, and food intake, the

route of human intake consists only of food and water intake. For human risk assessment with

regard to drinking water, daily consumption rates are compared to the acceptable daily intake

(ADI). ADI is a measure of the amount of a specific substance in food or drinking water that

can be ingested on a daily basis over a lifetime without an appreciable health risk. Murray et

al. (2010) proposed a list of priority MPs that are frequently detected and have an established

ADI value (Fig. 6).

More and more databases are being generated that provide information on toxic effects of

MPs. A database for pharmaceuticals currently includes data of 223 sources that studied 909

pharmaceuticals representing 79 classes of different drugs (www.wikipharma.org). In

addition, the PEIAR database for the preliminary risk assessment of common pharmaceuticals

for aquatic organisms has been generated (Cooper et al. 2008). MPs are also included in the

ECOTOX database of the US EPA, which contains toxic effects of 7,894 chemicals on 5,503

aquatic species and reported by 21,051 references as of 2011 (US EPA 2011b). Furthermore,

the ECOSAR program, freely available from US EPA can be used to predict acute and

chronic toxicity effects of chemicals on fish, aquatic invertebrates, and green algae (US EPA

2008b). In this case, the ecotoxicity is derived from known values for ‘similar’ compounds in

120 chemical classes by using over 440 (quantitative) structure activity relationships (see

also: Madden et al. 2009).

Fig. 6 Proposed priority list for MPs in drinking water (after Murray et al. 2010)

6.6 Laws and Regulations There are no specific regulations and acts concerning contamination by MPs, but MPs are

rather dealt with in conjunction with other pollutants. In the framework of the Clean Water

Act, the US EPA has developed a list of toxic pollutants (65 chemicals) to be regulated and

monitored, particularly in wastewater effluents, for protection of aquatic life in surface waters.

However, this list has changed little since its establishment in the mid-1970s. In addition,

there is a list of 126 priority pollutants of so far unregulated environmental contaminants that

industrials: PFOA, PFOS DEHP

pesticides: diazinon, methoxychlor, dieldrin

PPCPs: EE2, E1 carbamazepine, bE2, DEET, triclosan, acetaminophen

industrials: BDE-47, BDE-99

pesticides: benomyl, carbendazim, aldrin, endrin, ethion, malathion, biphenthrin,

cypermethrin

industrials: BPF

PPCPs: AHTN, HHCB, ibuprofen,

estriol

frequent occurrence ADI established

38

are to be monitored (US EPA http://water.epa.gov/scitech/methods/cwa/pollutants-

background.cfm). With regard to drinking water, the Chemical Candidate List is a list of 116

chemical compounds so far unregulated by US drinking water regulations that may need to be

regulated under the Safe Drinking Water Act. It includes pesticides, disinfection byproducts,

pharmaceuticals, and biological toxins (http://water.epa.gov/scitech/drinkingwater/dws/ccl/

ccl3.cfm).

In the European Union, the water framework directive established environmental quality

standards for surface waters for 33 priority substances plus 8 others (Directive 2008/105/EC,

http://ec.europa.eu/environment/water/water-framework/priority_substances.htm).

MPs are also covered by regulations and legislation concerning endocrine disruptors: the EU

has listed more than 550 substances, for which there is either clear evidence, or which are

suspected or considered to be endocrine disruptors (European Commission DG ENV 2007).

In the US, 207 chemicals and substances have been identified as priorities within the EPA’s

drinking water and pesticides programs (US EPA 2010b).

7. SELECTED EXAMPLES OF COMPOUNDS AND THEIR FATE IN THE

ENVIRONMENT 7.1 FATE OF ETHINYLESTRADIOL (EE2)

Ethinylestradiol is a synthetic hormone often used as an active substance in contraceptives.

The contraceptive pill is taken by 8.8 % of women in the reproductive age worldwide -

approximately 104 million women (Fig. 7). In actual formulations, the EE2 take up is about

20-35 µg per day. The human body excretes EE2 in form of metabolites – namely,

glucoronide and sulphate conjugates which are less active compounds. However, it is only by

chance that researchers detected that these compounds can be metabolized by microorganisms

in the sewage treatment plant to regenerate the ‘original’ EE2 with its full estrogenic potency

(Panter et al. 1999).

Fig. 7 Number of women of reproductive age taking contraceptive pill

(Source: United Nations 2011) EE2 can be found in wastewater and receiving rivers (Tab. 2). Often it is not the chemical

itself that is analytically determined, but rather the kind of hormonal activity that is measured

with bioassays. For example, the estrogenic activity is measured with the yeast-screen assay

(YES). This method is easier and less expensive than analytical chemistry. Moreover, it

makes sense since it is the biological activity that – eventually – is significant for organisms.

The determined biological activity is expressed as estradiol equivalents.

It is important to know that a great amount of other compounds show the same mode of action

as EE2, namely, binding at the estrogen receptor and are therefore called xenoestrogens.

When an effect due to estrogenic exposure is suspected, it helps to detect the estrogenic

activity in the environment with a bioassay, such as YES. Due to cost, identification of the

respective compound is often not performed. However, the aforementioned other

xenoestrogens are of considerably lower potencies. For example, bisphenol A is about 20,000

times less potent than E2 (Tab. 7). In addition, the potency also depends on the endpoint

measured. Apart from YES, several other bioassays can be used to elucidate the mode of

action by measuring the endpoints.

0 20 40 60 80 100 120

OCEANIA

NORTHERN AMERICA

LATIN AMERICA AND THE CARIBBEAN

EUROPE

ASIA

AFRICA

Less developed regions

More developed regions

WORLD

Mio women taking pill

40

Tab. 7 Estrogen equivalent factor (EEF) of hormones and chemicals in relation to 17ß-estradiol (adapted from Campbell et al. 2006; citations therein)

Compound EEF

17ß-estradiol (E2) 1.0

estrone (E1) 0.01 - 0.1

ethinylestradiol (EE2) 0.8 - 1.9

estriol (E3) 0.01 - 0.08

bisphenol A 5.0 x 10-5 - 6.0 x 10-5

nonylphenol (NP) 7.2 x 10-7 - 1.9 x 10-2 nonylphenol ethoxylates 2.0 x 10-7 - 1.3 x 10-5

octylphenol 1.0 x 10-5 - 4.9 x 10-4

When EE2 and other environmental chemicals with estrogenic-like activity enter the

watercourses, fish are among the most critically exposed group and also present the most

sensitive group for adverse effects of EDC. In addition, fish were most intensively studied.

Here, only a general insight can be provided: exposure of male fish to concentrations as

reported in many sewage plant effluents (0.1 ng/L) results in the synthesis of VTG – an egg-

yolk precursor protein that naturally only occurs in mature females under the influence of the

female sex steroid hormone estradiol. As well, retardation of the growth and development of

testes were stated in various fish species. In some species, the development of intersex was

noted, in which ova begin to be synthesized in testes (Walker et al. 2006). As a consequence,

reduced reproductive capabilities and, eventually, infertility is suspected. Indeed, lifelong

exposure to 5 ng EE2, leads to infertility in male zebrafish with subsequent complete

population failure (Nash et al. 2004). Interestingly, infertile male fish still showed normal

reproductive behaviour, and competed with fertile males, thereby reducing the fertilization

success of the latter.

While fish species are the most investigated taxa, other vertebrates also showed detrimental

reproductive disturbances at environmentally relevant concentrations. For example, exposure

of adult male frogs lead to sex reversal, disrupted reproductive organ development, and

reduced fertility, thereby indicating that estrogenic active compounds may pose a threat to

wild frog populations (Gyllenhammar et al. 2009).

It is difficult to prove that a specific compound is responsible for population decline in the

wild due to the long time delay and the multitude of additional interacting factors (Burkhardt-

Holm et al. 2005). As a result of the need for long time exposure as well as ethical

considerations of long-term animal experiments, only few experiments have been carried out

41

to demonstrate a population level effect of EE2. However, in the experimental lake area in

Canada, experimental concentrations of 5 ng/L EE2 lead to a feminization of male fathead

minnow after approximately one year and a total population collapse after another three years.

As consequence of the disappearance of this food source, there was also a 30% reduction of

lake trout (Halford 2008). Concentrations in the range found to impair reproduction in the

wild as well as in various species in laboratory experiments were detected in many water

bodies in almost every country investigated (Halford 2008).

In STPs, EE2 is biotransformed by nitrifying bacteria such as Nocardia sp. and ammonia-

oxidizing bacteria like Nitrosomonas spp. through oxidation followed by complete

mineralization to CO2 and water (Shi et al. 2004). Due to (bio)degradation, the chemical (or

biological) dissolution of materials by bacteria or other biological means, concentrations of

EE2 usually decrease in the order STP influent, STP effluent, surface water, ground water,

raw drinking water, and then drinking water (Adler et al 2001, Tab. 8). Concentrations of EE2

in drinking water were measured in several countries (e.g. Germany). Adler et al. (2001)

report a maximum concentration of 1.4 ng/L respectively 2.4 ng/L including conjugates. This

pollution may result from untreated or insufficiently treated groundwater containing EE2 in

which a reflux of pharmaceutical and synthetic estrogens from wastewater discharge remain.

The concentrations are not considered to be a risk (Adler et al. 2001).

Tab. 8 Fate of hormones in the environment, concentrations in ng/L (medians); values for free

EE2 (not-conjugated, ncon) and concentrations including conjugates (+con); <: below detection limit (adapted from Adler et al. 2001)

STP influent STP effluent surface water ground water raw drinking

water drinking water

ncon +con ncon +con ncon +con ncon +con ncon +con ncon +con

EE2 7 9.5 0.3 0.5 < < ? < < < < <

E2 1.5 3 0.2 0.5 < 0.1 < < < < < <

Estrone 5.5 13 2.5 8 < 0.4 < 0.1 < 0.2 < <

Another pathway for exposure of humans is via the food chain, when EE2 from vegetables,

milk, eggs or meat enter the human body (Stefanidou et al. 2009). In the US, the average daily

intake of endogeneous estrogens via food is estimated at 70 µg for males and 58 µg for

females (Caldwell et al. 2010).

The question remains whether there is a risk for humans or wildlife due to exposure of EE2 or

other estrogenically active compounds with the same mode of action. Accordingly, since once

42

a bioassay is used for measurement of endocrine activity, the specific MPs are obsolete; the

data and facts below refer to those specific compounds that have the same mode of action.

In general, there are several challenges making it difficult to assess the risk for humans and

wildlife:

(i) Effects are most often due to chronic exposure at low concentrations.

Especially for EE2, concentrations are at or below the limit of detection.

(ii) The long delay between exposure and effect makes it difficult to establish clear

cause-effect relationships. Exposure to xeno-estrogens during embryonic

development may lead to reproductive failure in adulthood. Even worse,

effects may be manifested only some generations later. These so-called

transgenerational effects are due to epigenetic events (e.g. alterations in the

DNA packing).

(iii) Humans are exposed to a natural exogenous burden by estrogens in food, such

as phytoestrogens. Although classified as weak estrogens, these hormones can

affect the endocrine system of humans: the daily consumption of 60g of

textured soy protein for one month, which is equivalent to 45mg isoflavones -

the estrogenic compound of soy - significantly decreased follicle stimulating

and luteinizing hormone levels and increased menstrual cycle length (Cassidy

1996).

(iv) Effects depend on sex and age and potential susceptible periods in

development. Sensitive time periods during development of the embryo exist

when hormones can set the course for the future with regard to organizational

changes, mostly irreversible effects; these are called sensitive windows.

Gestation weeks 7 to 24 in male foetuses, for example, represent such a

sensitive window. Spermatogenesis is negatively affected by various factors to

which the mother was exposed. Puberty in young boys is also a sensitive

window when spermatogenesis can be affected. Correspondingly, lack of

exposure data during critical periods of development are one of the biggest

hindrances for determining whether observed effects are linked to preceding

exposure to estrogenic compounds.

43

Apart from these general remarks, there is a wealth of studies raising concern that several

diseases and malformations of the reproductive tract could be due to exposure to

xenoestrogens in humans. These concerns are partly based on experimental studies with

vertebrates, especially mammals, and partly based on epidemiological studies.

Experimental studies with rodents showed that exposure to estrogenic disruptors during fetal

and prenatal stages induced premalignant and malignant transformation of the adult mammary

gland. For women, it is well known that estrogen exposure throughout a woman’s life is a

major risk factor for the development of breast cancer (Markey et al. 2003). Accordingly, it

has repeatedly been suggested that culmination of exposure to the multitude of environmental

hormones, together with medically used hormones, may cause an increase in breast cancer

during the last 50 years (1% per year) (Markey et al. 2003). This is the most frequent cancer

in women and the leading cause of cancer death worldwide (Fenichel and Brucker-Davis

2008). The positive correlation between increased intrauterine levels of estrogens and breast

cancer in daughters born from such pregnancies support this link (Stefanidou et al. 2009).

There are several more studies showing co-incidence between such impaired structures and

functions of the reproduction with exposure data. For example, high concentrations of

organochlorine pesticides (e.g. DDT, and its breakdown product DDE) in breast milk of

mothers are associated with cryptorchidism in boys (Damgaard et al. 2006). However, it is

difficult to establish a clear cause-effect relationship.

One well-documented issue is how exposure to several EDCs can disrupt masculinization.

Dioxin, for example, was found to result in lower sperm counts and reduced sperm motility in

adults who were exposed before puberty (Mocarelli et al. 2008). The exposure of parents to

dioxin during the Seveso accident in Italy in 1976 led to a significantly increased ratio of

newborn girls in comparison to boys (Mocarelli et al. 1999). Many studies were carried out in

mammals, suggesting that endocrine disruption leading to reproductive impairment could also

be the case in men (Sharpe 2010). However, clear cause-effect relationships as well as

extrapolation to humans are difficult to establish due to the aforementioned knowledge gaps.

Some studies suggest a relationship between the exposure to xenoestrogens and testicular

cancer – the most common cancer in young men. Incidences have almost doubled within the

last decades with high annual increases in low incidence countries (e.g. Spain) and low annual

increases in high incidence countries (e.g. Switzerland and Norway) (Bray et al. 2006).

44

Apart from dioxin, there is currently only minimal evidence indicating that environmental

exposure to EDCs is directly responsible for impairing the fertility in man (Sharpe 2010).

Other factors, such as obesity, sedentary work, and lifestyle may have higher potential to

adversely impact sperm production than environmental exposure to xenoestrogens (Sharpe

2010). However, exposure in occupational settings was shown to affect reproductive health:

men working in factories producing oral contraceptives developed gynecomastia and reported

a loss of libido (Degen and Bolt 2000); female workers experienced cycle irregularities

(Degen and Bolt 2000).

Taken together, dietary and occupational exposure to xenoestrogens and phytoestrogens were

shown to have an effect on humans. However, the mentioned difficulties make it impossible

to detect a clear cause-effect relationship between exposure to environmental xenoestrogens

and adverse human health (Degen and Bolt 2000).

7.2 FATE OF NONYLPHENOL (NP)

Nonylphenols (NPs) are synthetic organic compounds that do not occur naturally in the

environment. Nonylphenols and their derivatives have been banned in some products like

detergents by the OSPAR convention in 2000 and by the European Commission in 2005, but

are still permitted in other products in other regions of the world. They are problematic

because they are widespread, toxic, and persistent. They have been detected in wastewater

effluents, surface and groundwater, sediments, air, aquatic organisms, and food (Tab. 9).

Chronic risks have hardly been investigated, since only low concentrations have been

detected.

NPs are degradation products of nonylphenol ethoxylates (NPEOs), which are extensively

used as non-ionic surfactants in industrial chemicals (liquid soaps and cleaners) and

household products (liquid detergents), paints, cosmetics, or in biocides. The global

production of NPEOs is estimated at 500,000 tons per year of which 60% reach aquatic

environments (see Ying et al. 2002). Estimated NP and NPEO emissions into European

surface waters are 2,979 kg per day and 108,060 kg per day into wastewater (EU 2002). NPs

are released to the aquatic environment mainly via industrial and municipal STP effluents.

Leaking septic systems are minor sources and may directly discharge into the groundwater.

45

NP was reported to be photochemically degraded with a 10 to 15 hour half-life. On the other

hand, due to its hydrophobic nature, NP appears to be strongly adsorbed to sludge, soil, and

sediments especially when rich in organic carbon, clay, or silt where it is resistant to

biodegradation. When sludge is applied to agricultural fields, the NPs contaminate the soil.

Aerobic conditions favour biodegradation by microorganisms via stepwise loss of ethoxy

groups or carboxylation to alkylphenol ethoxycarboxylates (Fig. 8). Anaerobic conditions,

such as in sewers, sediments, and in some STP compartments favour degradation of NPOEs

to NP, which is resistant to biodegradation. In general, shorter-chain metabolites (mono-tri

ethoxylates, nonylphenol) may be more persistent and more toxic than the parent products.

Both aquatic and terrestrial plants take up small amounts of NP and transport it to other parts

such as shoots. Being lipophilic, NPs and NPEOs accumulate in aquatic organisms like plants,

algae, fish, and mussels with bioconcentration factors 1-410 for fish and up to 10000 for algae

(Ahel et al. 1993). Although NPs are transfered in the food chain, there are no reports about

biomagnification in the food chain. It is not yet clear if plants are able to metabolize and

degrade NP. However, the plant root releases exudates, which enhance microorganisms that

may be able to remove NP.

Fig. 8 Biodegradation of nonylphenol ethoxylate under aerobic and anaerobic conditions

(adapted from Porter and Hayden 2002) In animals, NP has a receptor-mediated mode of action and mimics the natural hormone 17-ß-

estradiol. Therefore, it may interact with the estrogen receptor and cause endocrine disruption.

However, its potency is much lower than that of the natural hormone: The EEF relative to 17-

Nonylphenol ethoxylate (NPEO)

NPEO with carboxylated

ethoxylate chain

short-chained NPEO

Nonylphenol

(NP)

aerobic )

anaerobic )

46

ß-estradiol is 7.2 x 10-7 - 1.9 x 10-2 for NP, and 2.0 x 10-7 - 1.3 x 10-5 for nonylphenol

ethoxylates (see also Tab. 7) (Folmar et al. 2002).

Nevertheless, NP was shown to inhibit soil organisms, induce VTG in fish, and enhance the

growth of human breast cancer cells (Tab 9). Acute toxicity of aquatic organisms was

reported at 17-3000 µg/L (Tab. 9).

Tab. 9 Endpoints and critical concentrations of nonylphenol (data from Servos 1999) Endpoint Effect concentration (µg/L)

NOEC invertebrates 3.7 NOEC fish 6

estrogenic in fish 10

acute toxicity to aquatic organisms 17-3000

algae growth 27 -bis 1500

LC50- / EC50 for fish 130 - 1400

LC50- / EC50 for invertebrates 180 - 3000

Almost all human food contains traces of NP, especially shellfish and seafood. In Taiwan, for

example, exposure was three times higher than in Germany (Tab. 10). However, toxicological

tests with rats could not prove that this uptake presents a risk for humans (Bontje et al. 2004).

Tab. 10 Human exposure to nonylphenol (adapted from Lu et al. 2007) country/area average daily intake

(µg/d)

Germany 7.5

Taiwan 28.0

New Zealand 3.3

8. POSSIBLE MEASURES

8.1 Source control In the long run, the most effective and socio-economically cheapest method may be source

control - reducing the production, use, and consumption of compounds. Although determining

substitutes often involves considerable cost, many products already have substitutes with less

harmful effects and others may follow. Here, the chemical and pharmaceutical industries have

an important role with regard to research and development. Some success has already been

made in the field of green chemistry, the “design, development, and implementation of

chemical products or processes to reduce or eliminate the use and generation of hazardous and

toxic substances” (Hjeresen et al. 2002). Pesticides such as DDT, lindane, aldrin, and dieldrin

47

have been replaced by carbamate and pyrethroid, which are less toxic and have a higher

biodegradability. Similar efforts should be made to substitute nonylphenol ethoxylates.

Source control may also involve a more conservative prescription practice by medical doctors

for instance with regard to synthetic hormones and antibiotics. ‘Green pharmacy’ aims at

developing drugs with reduced emissions into the environment (e.g. by preventing excretion

by humans), and improving degradability in water purification plants (Sumpter 2010). Careful

life cycle management such as ‘cradle-to-cradle design’ (Daughton 2003) comprises an

environmentally friendly manufacturing method, non-hazardous ingredients, and good

biodegradability. Further source control directed measures also include legislative measures,

restriction of use and changes in the consumer behaviour. Consumer acceptance is generally

increased when information campaigns are successful (Burkhardt-Holm and Götz 2011).

Accordingly, education and information are very valuable strategies in this context.

8.2 Separation of effluent streams

Separation of highly contaminated effluents, for example from hospitals or industry, is one

treatment measure that can lead to cost savings with regard to treatment. However, the effort

and costs associated with the necessary construction of new sewage systems and treatment

plants present a major hurdle in realizing this option. Nevertheless, in areas where sewer

systems have to be restored anyway, this may be a viable option.

8.3 Sewage Treatment In order to improve sewage treatment technology, ‘end-of-pipe’ efforts must be made to

ensure that concentrations of critical compounds in the effluent are constantly below the

PNEC (Schwarzenbach et al. 2006). One of the challenges is created by the fact that MPs

occur together with compounds with concentrations of a thousand to a million-fold higher

(Schwarzenbach et al. 2006). While conventional STPs are successful in removing or

transforming a number of MP substances, efficiency can be improved when experiments are

conducted and the treatment process is adapted to the specific sewage water. For example, the

removal efficiency depends on optimal sludge retention time, the adaptation of the community

of microorganisms, etc. A number of standard technologies for tertiary treatment of sewage

waste also can be used to increase removal of MPs from water or to transform them into

compounds of no hazard. Due to high costs, however, these methods are typically only used

for the treatment of drinking water. Although a complete removal of all pharmaceuticals,

48

pesticides, and industrial compounds will hardly be possible, the treatment of sewage waste

and the quality of the resulting water are important factors to be considered and – at present –

improved upon.

(i) by membrane bioreactors

In contrast to conventional biological treatment, membrane bioreactors include filtration through

micro- or ultrafiltration membranes that separate sludge and liquid (Cirja 2007). Clouzot et al.

(2010) indicate that membrane bioreactor technology may provide better steroid removal

compared to other treatment methods. Accordingly, membrane bioreactors could significantly

increase removal of the 1,6- and 2,7- naphthalene disulfonate (NDSA) and benzothiazole-2-

sulfonate (De Wever et al. 2007). Membrane filtration (e.g. nanofiltration) is also a technical

option being applied in drinking water production, where pesticides are expected in the raw

water, as well as in direct water reuse schemes (e.g. the Newater scheme in Singapore). Here,

high costs as well as energy requirements again present a barrier to the implementation of this

technology. Thus, it is expected that ozonation or activated carbon filtration are generally

more cost effective. (ii) by ozonation

Ozonation presents a viable option for the effective removal of several MPs (Ternes et al.

2003). Ozone (O3) is a reactive gas that when in contact with water forms radicals, breaks

down complex compounds, and makes them much more susceptible to biological degradation.

Good removal was shown, for example, for the anticonvulsant drug carbamazepine and the

anti-inflammatory drug diclofenac, naproxen and ibuprofen reducing levels between 70 and

100 per cent (Schaar et al. 2010, Pal et al. 2010). However, some compounds like

carbamazepine and diclofenac may be transformed into potentially harmful products (Escher

and Fenner 2011). Furthermore, the ozonation process also destroys microorganisms which

may be important for further degradation processes. A solution to this issue may be to include

further methods, such as sand filters with biofilms, following the ozonation process. In

Switzerland, the cost of ozonation (investment and operation) were estimated to range from

0.05 to 0.20 € m-3, respectively 5-20 Euro capita-1 a-1, depending on the size of the STP and

the contaminant load of the waste water (Joss et al. 2008).

(iii) by activated carbon

49

Activated carbon – carbon that has been processed to have a large surface area available for

adsorption or chemical reactions – has shown a high efficiency in removing many MPs,

especially those of a high hydrophobicity (Nowotny et al. 2007, Snyder et al. 2007). In

Switzerland, the cost of carbon adsorption were estimated to range from 0.08 to 0.20 € m-3

(Joss et al. 2008).

Experts suggest a combined treatment, comprising activated carbon and microfiltration

processes including reverse osmosis and nanofiltration, are showing the best results.

Unfortunately, combined treatments are often too expensive in practice. In the end, cost

calculations must be made in conjunction with the goals of water treatment; thus, the

prospective use of the treated water proves to be a decisive factor.

9. CONCLUSION AND OUTLOOK

Even at small concentrations, exposure to environmental MPs was shown to affect wildlife

already. Despite the number of standardized toxicity tests and models available, relationships

between concentrations, exposure, short-term effects, long-term impacts, and effects on the

population and species levels are still poorly understood (Pal et al. 2010). Thus far, significant

adverse effects on human health have not been proven (Degen and Bolt 2000).

However, with regard to the uncertainties, it seems advisable to apply the precautionary

principle to any risk assessment. “The precautionary principle provides justification for

public policy action in situations of scientific complexity, uncertainty and ignorance, where

there may be a need to act in order to avoid or reduce potentially serious or irreversible

threats to health or the environment, using appropriate level of scientific evidence, and taking

into account the likely pros and cons of action and inaction” (Gee 2006).

Furthermore, due to the population increase and the increasing average life expectancy,

greater access to health care, and new drug developments, the input of pharmaceutical

residues into the aquatic environment is likely to increase. Finally, increasing water scarcity

resulting from climate change, higher water consumption by households, and increasing

withdrawal of water for industrial and agricultural purposes will eventually lead to a

concentration of the pollutants in the remaining surface and ground water if they are not

counterbalanced by new product engineering or treatment technologies.

50

Such treatment technologies include source control as well as ‘end-of-the-pipe’ measures.

Source control, such as restricted use of MPs, green pharmacy, and green chemistry, are

efficient and represent the best solution. However, source control often requires a high input

of money and know-how as well as societal change. For this reason, tertiary wastewater

treatment technologies that are able to remove a significant level of MPs from surface or

drinking water resources are typically the method of choice in reducing risks for consumers

and the environment.

As the most important research needs and actions, we suggest:

• more detailed fate studies that provide information on photo- and biotransformation,

abiotic transformation, as well as sorption and desorption processes in dependence of

travel time and distance; these studies should include metabolites and conjugates and

should be carried out in climate zones where they are largely lacking (e.g Indonesia,

India, Canada, Russia and countries from Africa and South America) (Pal et al. 2010);

• toxicity tests that include a wider number of marine and freshwater species and address

metabolites and mixtures;

• improvement of the understanding of dose-effect relationships;

• research regarding the relevance of micropollutants for the environment as a whole; this

research should include long term studies as well as experiments under real world

conditions, such as fluctuating concentrations and seasonal dynamics;

• further development of models that can help extrapolate knowledge on effects to other

species and, thereby serve to improve and expand risk assessment;

• integration of knowledge into freely accessible databases;

• inclusion of the removability of MPs by advanced sewage treatment in registration tests

required for registration of chemical compounds (Joss et al 2007);

• mitigation measures aimed at reducing the input of MPs into water systems

• development and application of measures for success control.

51

Acknowledgement

This paper would not have been possible without the help of Susanne Wolfer, Tara Hadler,

and Karla Schlie.Thanks to Susanne for her help with the general paper – in particular with

the literature review. Thanks also to Tara for her help in editing this paper and Karla for her

work with most of the figures.

52

Glossary of Terms (adapted from http://www.atsdr.cdc.gov/glossary.html)

Absorption

Process in which molecules present in a given fluid enter into another bulk phase

Adsorption

Physical adherence onto the surface of another molecule or substance

Acceptable daily intake

A measure of the amount of a specific substance in food or water that can be ingested on a daily basis

over a lifetime without an appreciable risk

Acute toxicity

Toxicity due to contact with a substance that occurs for only a short time (up to 14 days)

Additive effect

A biologic response to exposure to multiple substances that equals the sum of responses of all the

individual substances added together

Aerobic

Requiring oxygen

Anaerobic

Requiring the absence of oxygen

Attenuation

Reduction in concentration; natural attenuation of chemicals in rivers is obtained by the so-called self-

cleaning capacity

Biodegradation

Decomposition or breakdown of a substance through the action of microorganisms (such as bacteria or

fungi)

Body burden

The total amount of a substance in the body. Some substances build up in the body because they are

stored in fat or bone or because they leave the body very slowly

Carcinogen

A substance that causes cancer

Chronic toxicity

Toxicity due to contact with a substance that occurs over a long time

Contaminant

A substance that is either present in an environment where it does not belong or is present at levels

that might cause harmful (adverse) health effects

53

Contaminant of emerging concern (CEC)

See emerging (environmental) contaminant

Degradation

Process of a conversion or breakdown of a substance from one form to another by physical, chemical

or biological means

Derivative

A compound derived or obtained from another and containing essential elements of the parent

substance

Dose-response relationship

The relationship between the amount of exposure (dose) to a substance and the resulting changes in

body function or health (response)

Emerging (environmental) contaminant

Chemicals that have only recently been analysed or identified in the environment and which are

believed to cause adverse effects on ecosystems and humans

Endocrine

Pertaining to hormones or to the glands that secrete hormones

Endpoint

Measurable parameter that indicates a preceding exposure or the effect of a chemical; it constitutes one

of the target observations of the trial

E-Screen

Laboratory test that measures growth of MCF-7 cells in vitro in response to endocrine substances.

MCF-7 is a breast cancer cell line, the acronym refers to the institute where the cell line was

established.

Exposure

Contact with a substance. Exposure may be short-term (acute exposure), of intermediate duration, or

long-term (chronic exposure)

Gynecomastia

The abnormal development of large mammary glands in males resulting in breast enlargement

Half-life

The time it takes for half the original amount of a substance to disappear, either due to elimination or

due to degradation

Hazard

A source of potential harm from past, current, or future exposures

In vitro

54

In an artificial environment outside a living organism or body. For example, some toxicity testing is

done on cell cultures or slices of tissue grown in the laboratory, rather than on a living animal

In vivo

Within a living organism or body; for example, some toxicity testing is done on whole animals, such

as rats or mice

Lowest-observed-effect concentration (LOEC)

See LOEL, the lowest tested dose of a substance that has been reported to cause harmful (adverse)

effects in organisms

Lowest-observed-effect level (LOEL)

The lowest concentration or amount of a substance found by experiment or observation which causes

an adverse alteration of morphology, function, capacity, growth, development or life span of a target

organism distinguished from normal organisms of the same species under defined conditions of

exposure

Metabolism

The conversion or breakdown of a substance from one form to another by a living organism

Metabolite

Any product of metabolism

Mutagen

A substance that causes genetic damage

No-observed-effect concentration (NOEC)

The highest tested dose of a substance that has been reported to have no harmful (adverse) health

effects

No-observed-effect level (NOEL)

See NOEC, an exposure level at which there are no statistically or biologically significant increases in

the frequency or severity of adverse effects between the exposed population and its appropriate

control; some effects may be produced at this level, but they are not considered as adverse, or as

precursors to adverse effects

PEC

Predicted environmental concentration; derived from consumption studies and/or models

PNEC

Predicted no effect concentration; usually derived by dividing LOEC or NOEC by safety factor

Quantitative structure-activity relationship (QSAR)

55

The process by which chemical structure is quantitatively correlated with a well defined process, such

as biological activity or chemical reactivity

Safety factor

Mathematical adjustments for reasons of safety when knowledge is incomplete. Uncertainty factors are

used to account for variations in sensitivity of individuals of the same species, for differences between

test species and target species, for differences between animals and humans, and for differences

between a LOEC and a NOEC.

Synergistic effect

A biologic response to multiple substances where one substance worsens the effect of another

substance. The combined effect of the substances acting together is greater than the sum of the effects

of the substances acting by themselves.

Tertiary treatment

Wastewater treatment following the biological secondary treatment, e.g. oxidation, ozonation

Secondary treatment

Wastewater treatment designed to substantially degrade the organic content of the sewage using

biological methods, e.g. activated sludge

Toxicity threshold

The exposure level or dose of an agent above which toxicity or adverse effects can occur, and below

which toxicity or adverse effects are unlikely.

Toxicological study

The study of how much poison must be present to produce an effect on animals or plant systems, may

also include what type of effect is produced and how it is detected

Vascular system

Plant tissues for conducting water, minerals, and photosynthetic products through the plant

Xenobiotic

Chemical compound foreign to the body or to living organisms such as a pesticide

Xenoestrogen

Synthetic molecule or exogenous natural estrogen that was taken up from the environment and mimics

the endogenous estrogen, i.e. behaves like a molecule synthesized by the body itself. Sometimes, only

synthetic chemicals with estrogenic properties are termed xenoestrogens, while natural estrogens (such

as those excreted by other animals or humans) as well as phytoestrogens and mycoestrogens

(synthesized by fungi and plants, respectively) are not included.

YES-Yeast Estrogen Screen

In vitro assay for xenoestrogens using yeast cells

56

References

Adler, P., Steger-Hartmann, T., and Kalbfus, W. (2001) Distribution of natural and synthetic steroid hormones in water samples from Southern and Middle Germany. Acta Hydrochim. Hydrobiol. 29: 227-241.

Ahel, M., McEvoy, J., and Giger, W. (1993) Bioaccumulation of the lipophilic metabolites on nonionic surfactants in freshwater organisms. Environ. Pollut. 79: 243-248.

Arikan, O.A., Rice, C., and Codling, E. (2008) Occurrence of antibiotics and hormones in a major agricultural watershed. Desalination 226: 121-133.

Beausse, J. (2004) Selected drugs in solid matrices: A review of environmental occurrence, determination and properties of principal substances. Horizontal 26: 1-36. http://www.ecn.nl/docs/society/horizontal/hor_desk_26_pharmaceuticals.pdf

Besse, J.P., Kausch, Barreto, C., and Garric, J. (2008) Exposure assessment of pharmaceuticals and their metabolites in the aquatic environment: application to the French situation and preliminary prioritization. J. Hum. Ecol. Risk Assess. 14: 665-695.

Bokern, M., Raid, P., and Harms, H. (1998) Toxicity, uptake and metabolism of 4-n-nonylphenol in root cultures and intact plants under septic and aseptic conditions. Environ. Sci. Pollut. Res. 5: 21-27.

Bontje, D., Hermens, J., Vermeire, T., and Damstra, T. (2004) Integrated Risk Assessment: Nonylphenol Case Study. Report Prepared For The Who/Unep/Ilo International Programme On Chemical Safety. WHO/IPCS/IRA/12/04. http://www.who.int/ipcs/methods/Nonylphenol.pdf

Boxall, A.B.A., Sinclair, C.J., Fenner, K., Kolpin, D., and Maud, S.J. (2004) When synthetic chemicals degrade in the environment. Environ. Sci. Technol. 38: 368A-375A.

Bray, F., Richiardi, L., Ekbom, A., Pukkala, E., Cuninkova, M., and Moller, H. (2006) Trends in testicular cancer incidence and mortality in 22 European countries: continuing increases in incidence and declines in mortality. Int. J. Cancer 118: 3099-111.

Brian, J.V., Harris, C.A., Scholze, M., Backhaus, T., Booy, P., Lamoree, M., Pojana, G., Jonkers, N., et al. (2005) Accurate prediction of the response of freshwater fish to a mixture of estrogenic chemicals. Environ. Health Perspect. 113: 721-728.

Briggs, G.G, Bromilow, R.H., Evans, A.A., and Williams, M. (1983) Relationships between lipophilicity and the distribution of non-ionized chemicals in barley shoots following uptake by the roots. Pestic. Sci. 14: 492-500.

Bucheli, T.D., Wettstein, F.E., Hartmann, N., Erbs, M., Vogelgsang, S., Forrer, H.R., and Schwarzenbach, R.P. (2008) Fusarium mycotoxins: Overlooked aquatic micropollutants? J. Agric. Food Chem. 56: 1029-1034.

Burkhardt-Holm, P., Giger, W., Güttinger, H., Ochsenbein, U., Peter, A., Scheurer, K., Segner, H., et al. (2005) Where have all the fish gone? Environ. Sci. Technol. 39: 441A - 447A.

Burkhardt-Holm, P., and Götz, K. (2011) Environmental hormones: risk, fate and potential measures to reduce their entry into the water cycle. Poster presented at the Singapore International Water Week 2011, July 04-08. Singapore.

Caldwell, D.J., Mastrocco, F., Nowak, E., Johnston, J., Yekel, H., Pfeiffer, D., et al. (2010) An Assessment of Potential Exposure and Risk from Estrogens in Drinking Water. Environ. Health Perspect. 118: 338-344.

Campbell, C.G., Borglin, S.E., Bailey Green , F., Grayson, A., Wozei, E., and Stringfellow, W.T. (2006) Biologically directed environmental monitoring, fate, and transport of estrogenic endocrine disrupting compounds in water: A review. Chemosphere 65: 1265-1280.

Carballa, M., Omil, F., Lema, J.M., Llompart, M., Garcia-Jares, C., Rodriguez, I., Gomez, M., and Ternes, T. (2004) Behavior of pharmaceuticals, cosmetics andhormones in a sewage treatment plant Water Res. 38: 2918-2926.

57

Carr, D.L., Morse, A.N., Zak, J.C., and Anderson, T.A. (2011) Microbially mediated degradation of common pharmaceuticals and personal care products in soil under aerobic and reduced oxygen conditions. Water Air Soil Pollut. 216: 633-642.

Cassidy, A. (1996) Physiological effects of phyto-estrogens in relation to cancer and other human health risks. Proc. Nutr. Soc. 55: 399-417.

Cirja, M. (2007) Studies on the Behaviour of Endocrine Disrupting Compounds in a Membrane Bioreactor. Dissertation, RWTH Aachen. http://darwin.bth.rwth-aachen.de/opus3/volltexte/2008/2180/pdf/Cirja_Magdalena.pdf

Clara, M., Gans, O., Windhofer, G., Krenn, U., Hartl, W., Braun, K., Scharf, S., and Scheffknecht , C. (2011) Occurrence of polycyclic musks in wastewater and receiving water bodies and fate during wastewater treatment. Chemosphere 82: 1116-1123.

Clouzot, L., Doumenq, P., Vanloot, P., Rochea, N., and Marrot, B. (2010) Membrane bioreactors for 17α-ethinylestradiol removal. J. Membrane Sci. 362: 81-85.

Cook, P.M., Erickson, R.J., Spehar, R.L., Bradbury S.P., and Ankley, G.T. (1993) Interim report on data and methods for assessment of 2,3,7,8-tetrachlorodibenzo-p-dioxin risks to aquatic life and associated wildlife. EPA-600/R-93/055. U.S. Environmental Protection Agency, Duluth, MN.

Cooper, E.R., Siewicki, T.C., and Phillips, K. (2008) Preliminary risk assessment database and risk ranking of pharmaceuticals in the environment. Sci Total Environ. 398: 26-33.

Damgaard, I.N., Skakkebaek, N.E., Toppari, J., Virtanen, H.E., Shen, H.Q., Schramm, K.W., Holm, P.J., Kold, J.T., and Main, K.M. (2006) Persistent pesticides in human breast milk and cryptorchidism. Environ. Health Perspect. 114: 1133-1138.

Daughton, C.G. (2003) Cradle-to-Cradle Stewardship of Drugs for Minimizing Their Environmental Disposition While Promoting Human Health. I. Rationale for and Avenues toward a Green Pharmacy. Environ. Health Perspect. 111: 757-774.

Degen, G.H., and Bolt, H.M. (2000) Endocrine disruptors: update on xenoestrogens. Int. Arch. Occup. Environ. Health 73: 433-441.

Della Greca, M., Pinto, G., Pistillo. P., Pollio, A., Previtera, L., and Temussi, F. (2008) Biotransformation of ethinylestradiol by microalgae. Chemosphere 70: 2047-2053.

De Wever, H., Weiss, S., Reemtsma, T., Vereeckena, J., Muller, J., Knepper, T., Rordend, O., Gonzalez, S., et al. (2007) Comparison of sulfonated and other micropollutants removal in membrane bioreactor and conventional wastewater treatment. Water Res. 41: 935-945.

Dewi R., R., and Pertiwi, G. (2006) Monitoring of Banned Pesticides in Indonesia. International POPs Elimination Project - IPEP. http://www.ipen.org/ipepweb1/library/ipep_pdf_reports/ 3ins%20monitoring%20banned%20pesticides%20in%20indonesia.pdf

Ellis, J.B. (2006) Pharmaceutical and personal care products (PPCPs) in urban receiving waters Environ. Pollut. 144: 184-189.

EMEA (2006) European Medicine Agency guideline on the environmental risk assessment of medicinal products for human use. EMEA/CHMP/SWP/4447/00.

Endo, S., Takizawa, R., Okuda, K., Takada, H., Chiba, K., Kanehiro, H., Ogi, H., Yamashita, R., and Date, T. (2005) Concentration of polychlorinated biphenyls (PCBs) in beached resin pellets: variability among individual particles and regional differences. Marine Pollut. Bull. 50: 1103-1114.

Engraff, M., Solere, C., Smith, K.E.C., Mayer, P., and Dahllöf, I. (2011) Aquatic toxicity of PAHs and PAH mixtures at saturation to benthic amphipods: Linking toxic effects to chemical activity. Aquat. Toxicol. 102: 142-149.

Escher, B.I., and Fenner, K. (2011) Recent Advances in Environmental Risk Assessment of Transformation Products. Environ. Sci. Technol. 45: 3835-3847.

58

Escher, B.I., and Hermens, J.L.M. (2004) Internal Exposure: Linking Bioavailability to Effects. Environ. Sci. Technol. 38: 455A-462A.

EU (2002) Continental Emissions. http://www.umweltdaten.de/wasser/themen/stoffhaushalt/nonylphenol.pdf.

European Commission DG ENV (2007) Study on enhancing the endocrine disrupter priority list with a focus on low productivity volume chemicals. Ref. ENV.D.4/2005/0028r. http://ec.europa.eu/environment/endocrine/documents/final_report_2007.pdf

Fahlenkamp, H., Nöthe, T., Nowotny, N., Ries, T., Hannich, C.B., Peulen, C. et al. (2006). Untersuchungen zum Eintrag und zur Elimination von gefährlichen Stoffen in kommunalen Kläranlagen, Teil 2.- MUNLV – NRW 1.16.100.1.

Fenichel, P., and Brucker-Davis, F. (2008) Breast risk cancer and environmental endocrine disruptors. Gynecologie Obstetrique et Fertilite 36: 969-977.

Fick, J., Soderstrom, H., Lindberg, R.H., Phan, C., Tykslind, M., and Larsson, D.G.J. (2009) Contamination of surface, ground, and drinking water from pharmaceutical production. Environmental Toxicology and Chemistry 28: 2522-2527.

Freidig, A.P., Verhaar, H.J.M., and Hermens, J.L.M. (1999) Comparing the potency of chemicals with multiple modes of action in aquatic toxicology: acute toxicity due to narcosis versus reactive toxicity of acrylic compounds. Environ. Sci. Technol. 33: 3038-3043.

Folmar, L.C., Hemmer, M.J., Denslow, N.D., Kroll, K., Chen, J., Cheek, A., Richman, H., Meredith, H., and Grau, E.G. (2002) A comparison of the estrogenic potencies of estradiol, ethynylestradiol, diethylstilbestrol, nonylphenol and methoxychlor in vivo and in vitro. Aquat. Toxicol. 60: 101-110.

Gee, D. (2006) Late lessons from early warnings: toward realism and precaution with endocrine disrupting substances. Environ. Health Perspect. 114(Suppl. 1): 152-160.

Gyllenhammar, I., Holm, L., Eklund, R., and Berg, C.(2009) Reproductive toxicity in Xenopus

tropicalis after developmental exposure to environmental concentrations of ethynylestradiol. Aquat. Toxicol. 91: 171-178.

Halford, B. (2008) Side effects. Chem. Eng. News 86: 13-17.

Helbling, D.E., Hollender, J., Kohler, H.E., Singer, H., and Fenner, K. (2010) High-throughput identification of microbial transformation products of organic micropollutants. Environ. Sci. Technol. 44: 6621-6627.

Herklotz, P.A., Gurung, P., Vanden Heuvel, P., and Kinney, C.A. (2010) Uptake of human pharmaceuticals by plants grown under hydroponic conditions. Chemosphere 78: 1416-1421.

Hjeresen, D.L., Kirchhoff, M.M., and Lankey, R.L. (2002) Green Chemistry: Environment, Economics, and Competitiveness. Corporate Environ. Strategy 9: 259-266.

Hook, S.E., and Fisher, N. S. (1998) Silver accumulation and toxicity in marine and freshwater zooplankton. In: Andern, A.W. (ed.) 5th International Conference on Transport Fate and Effects of Silver in the Environment, Hamilton, Ontario, September 28-October 1, 1997: 265-273.

IMS (2010) IMS Health Market Prognosis, March 2010. www.imshealth.com

Joss, A., Siegrist, H., and Ternes, T. (2007) Priority Compounds and Water Reuse. Neptune Workshop 2. http://www.wrrs2007.org/content/user/File/Workshop%202%20PRIORITY%20POLLUTANTS% 20extended%20Abstract.pdf

Joss, A., Siegrist, H., and Ternes, T.A. (2008) Are we about to upgrade wastewater treatment for removing organic micropollutants? Water Sci. Technol. 57: 251-255.

59

Kameda, Y., Tamada, M., Kanai, Y. , and Masunaga, S. (2007) Occurrence of Organic UV Filters in Surface Waters, Sediments, and Core Sediments in Tokyo Bay, -Organic UV Filters are New POPs? Organohalogen Comp. 69: 253-266.

Kline (2010) Impact of Recessions on the Personal Care Market. http://www.in-cosmetics.com/files/mktg_impact_of_recessions_on_the_personal_care_market_kline.pdf

Kupper, T., Plagellat, C., Brändli, R.C., de Alencastro, L.F., Grandjean, D., and Tarradellas, J. (2006) Fate and removal of polycyclic musks, UV filters and biocides during wastewater treatment. Water Res. 40: 2603–2612.

Lin, A.Y.C., Yu, T.H., and Lin, C.F. (2008) Pharmaceutical contamination in residential, industrial, and agricultural waste streams: risk to aqueous environments in Taiwan. Chemosphere 74:131-141.

Liu, C.W., Lin, C.C., Jang, C.S., Sheu, G.R., and Tsui, L. (2009) Arsenic accumulation by rice grown in soil treated with roxarsone. J. Plant Nutr. Soil Sci. 172: 550-556.

Lu, Y.Y., Chen, M.L., Sung, F.C., Wang, P.S., and Mao, I.F. (2007) Daily intake of 4-nonylphenol in Taiwanese. Environ. Int. 33: 903-910.

Madden, J.C., Enoch, S.J., Hewitt, M., and Cronin, M.T.D. (2009) Pharmaceuticals in the environment: Good practice in predicting acute ecotoxicological effects. Toxicol. Lett. 185: 85-101.

Markey, C.M., Rubin, B.S., Soto, A.M., and Sonnenschein, C. (2003) Endocrine disruptors: from Wingspread to environmental developmental biology. J. Steroid Biochem. Mol. Biol. 83: 235-244.

Martin, O.V., and Voulvoulis, N. (2009) Sustainable risk management of emerging contaminants in municipal wastewaters. Phil. Trans. R. Soc. A 367: 3895-3922.

Meng, L., Qiao, M., and Arp, H.P.H. (2011) Phytoremediation efficiency of a PAH-contaminated industrial soil using ryegrass, white clover, and celery as mono- and mixed cultures. J. Soils Sed. 11: 482-490.

Mocarelli, P., Gerthoux, P.M., Brambilla, P., Marocchi, A., Beretta, C., Bertona, M., Cazzaniga, M., et al. (1999) Dioxin health effects on humans twenty years after Seveso: a summary. Chemistry, man and environment: the Seveso accident 20 years on - monitoring, epidemiology and remediation: 41-51.

Mocarelli, P., Gerthoux, P.M., Patterson, D.G., Milani, S., Limonta, G., Bertona, M., Signorini, S., Tramacere, P.L., Colombo, L., et al. (2008) Dioxin exposure, from infancy through puberty, produces endocrine disruption and affects human semen quality. Environ. Health Perspect. 116: 70-77.

Molander, L., Agerstrand, M., and Ruden, C. (2009) WikiPharma – A freely available, easily accessible, interactive and comprehensive database for environmental effect data for pharmaceuticals. Regul. Toxicol. Pharmacol. 55: 367-371.

Murray, K.E., Thomas S.M., and Bodour, A.A. (2010) Prioritizing research for trace pollutants and emerging contaminants in the freshwater environment. Environ. Pollut. 158: 3462-3471.

Nash, J.P., Kime, D.E., Van der Ven, L.T.M., Wester, P.W., Brion, F., Maack, G., Stahlschmidt-Allner, P., and Tyler, C.R. (2004) Long-term exposure to environmental concentrations of the pharmaceutical ethynylestradiol causes reproductive failure in fish. Environ. Health Perspect. 112: 1725-1733.

Newman, M.C., and Unger, M.A. (2003) Fundamentals of Ecotoxicology. Lewis Publishers, Boca Raton Fl. 458 p.

Nowotny, N., Epp, B., von Sonntag, C., and Fahlenkamp, H. (2007) Quantification and modeling of the elimination behavior of ecologically problematic wastewater micropollutants by adsorption on powdered and granulated activated carbon. Environ. Sci. Technol. 41: 2050-2055.

Osano, O., Admiraal, W., Klamer, H.J.C., Pastor, D., and Bleeker, E.A.J. (2002) Comparative toxic and genotoxic effects of chloroacetanilides, formamidines and their degradation products on Vibrio fischeri and Chironomus riparius. Environ. Pollut. 119: 195-202.

60

Overmyer, J.P., Armbrust, K.L., and Noblet, R. (2003) Susceptibility of black fly larvae (Diptera: Simuliidae) to lawn-care insecticides individually and as mixtures. Environ. Toxicol. Chem. 22: 1582-1588.

Pal, A., Gin, K.Y.H., Lin, A.Y.C., and Reinhard, M. (2010) Impacts of emerging organic contaminants on freshwater resources: Review of recent occurrences, sources, fate and effects. Sci. Total Environ. 408: 6062-6069.

Panter, G.H., Thompson, R.S., Beresford, N., and Sumpter, J.P. (1999) Transformation of a non-oestrogenic steroid metabolite to an oestrogenically active substance by minimal bacterial activity. Chemosphere 38: 3579-3596.

Passuello, A., Mari, M., Nadal, M., Schuhmacher, M., and Domingo, J.L. (2010) POP accumulation in the food chain: Integrated risk model for sewage sludge application in agricultural soils. Environ. Int. 36: 577-583.

Pomati, F., Castiglioni, S., Zuccato, E., Fanelli, R., Vigetti, D., Rossetti, C., et al. (2006) Effects of a complex mixture of therapeutic drugs at environmental levels on human embryonic cells. Environ. Sci. Technol. 40: 2442–2447.

Porter, A.J., and Hayden, N.J. (2002) Nonylphenol in the Environment: A Critical Review. http://www.emba.uvm.edu/~nhayden/npreview.pdf

Sanderson, H., Brain, R.A., Johnson, D.J., Wilson, C.J., and Solomon, K.R. (2004) Toxicity classification and evaluation of four pharmaceuticals classes: antibiotics, antineoplastics, cardiovascular, and sex hormones. Toxicology 203: 27-40.

Schaar, H., Clara, M., Gans, O., and Kreuzinger, N. (2010) Micropollutant removal during biological wastewater treatment and a subsequent ozonation step. Environ. Pollut. 158: 1399-1404.

Scharf, S., Gans, O., Sattelberger, R. (2002) Arzneimittelwirkstoffe im Zu- und Ablauf von Kläranlagen. Umweltbundesamt. Report BE-201.

Schmidt, K., Steinberg, C.E.W., Staaks, G.B.O., and Pflugmacher, S. (2005) Influence of a xenobiotic mixture (PCB and TBT) compared to single substances on swimming behavior or reproduction of Daphnia magna. Acta Hydrochim. Hydrobiol. 33: 287-300.

Schnell, S., Bols, N.C., Barata, C., and Porte, C. (2009) Single and combined toxicity of pharmaceuticals and personal care products (PPCPs) on the rainbow trout liver cell line RTL-W1. Aquat. Toxicol. 93: 244-252.

Schoknecht, U., Gruycheva, J., Mathies, H., Bergmann, H., and Burkhardt, M. (2009) Leaching of Biocides Used in Facade Coatings under laboratory Test Conditions. Environ. Sci. Technol. 43: 9321-9328.

Schwarzenbach, R.P., Gschwend, P.M., and Imboden, D.M. (2003) Environmental Organic Chemistry. 2nd ed., Wiley, Hoboken , NJ.

Schwarzenbach, R.P., Escher, B.I., Fenner, K., Hofstetter, T.B., Johnson, C.A., von Gunten, U., and Wehrli, B. (2006) The Challenge of Micropollutants in Aquatic Systems. Science 313: 1072-1077.

Servos, M. (1999) Review of aquatic toxicity, estrogenic responses and bioaccumulation of alkylphenols and alkylphenol polyethoxylates. Water Qual. Res. J. Canada 34: 123-177.

Sharpe, R.M. (2010) Environmental/lifestyle effects on spermatogenesis. Phil. Trans. Royal Soc. London B 365: 1697-1712.

Shi, W., Wang, L., Rousseau, D.P.L., and Lens, P.N.L. (2010) Removal of estrone, 17α-ethinylestradiol, and 17ß-estradiol in algae and duckweed-based wastewater treatment systems. Environ. Sci. Pollut. Res. 17: 824–833.

Skarphedinsdottir, H., Gunnarsson , K., Gudmundsson, G.A., and Nfon, E. (2010) Bioaccumulation and Biomagnification of Organochlorines in a Marine Food Web at a Pristine Site in Iceland. Arch. Environ. Contam. Toxicol. 58: 800–809.

61

Snyder, S.A., Westerhoff, P., Yoon, Y., and Sedlak, D.L. (2003) Pharmaceuticals, Personal Care Products, and Endocrine Disruptors in Water: Implications for the Water Industry. Environ. Eng. Sci. 20: 449-469.

Snyder, S.A., Adham, S., Redding, AM, Cannon, Fred S., DeCarolis, J., Oppenheimer, J., Wert, E.C., Yoon, Y. (2007) Role of membranes and activated carbon in the removal of endocrine disruptors and pharmaceuticals. Desalination 202: 156-181.

Snyder, S.A (2008) Occurrence, treatment, and toxicological relevance of EDCs and pharmaceuticals in water. Ozone: Science and Engineering 30: 65-69.

Snyder, S.A., Trenholm, R.A., Snyder, E.M., Bruce, G.M., Pleus, R.C., and Hemming, J.D.C. (2008) Toxicological Relevance of EDCs and Pharmaceuticals in Drinking Water. IWA Publishing, Denver, CO.

Stefanidou, M., Maravelias, C., and Spiliopoulou, C. (2009) Human exposure to endocrine disruptors and breast milk. Endocr. Metab. Immune Disord. Drug Targets 9: 269-276.

Sumpter, J.P. (2010) Pharmaceuticals in the Environment: Moving from a Problem to a Solution. In: Kümmerer, K. and Hempel, M. (eds): Green and Sustainable Pharmacy. Springer-Verlag Berlin, Heidelberg: 11-21.

Syracuse SRC Interactive PhysProp Database (2011) http://www.syrres.com/ (accessed May 24, 2010)

Tan Yong Soon, Lee Tung Jean and Karen Tan (2009). Clean, Green, and Blue : Singapore’s Journey Towards Environmental and Water Sustainability. Institute of Southeast Asian Studies, Singapore.

Teijon, G., Candela, L., Tamoh, K., Molina-Díaz, A., and Fernández-Alba, A.R. (2010) Occurrence of emerging contaminants, priority substances (2008/105/CE) and heavy metals in treated wastewater and groundwater at Depurbaix facility (Barcelona, Spain). Sci. Total Environ. 408: 3584-3595.

Ternes, T.A., Knacker, T., and Oehlmann, J. (2003) Koerperpflegemittel in der aquatischen Umwelt. Eine bisher vernachlaessigte Stoffgruppe. UWSF – Z. Umweltchem. Okotox. 15: 169-180.

Ternes, T.A., Joss, A., and Siegrist, H. (2004) Scrutinizing pharmaceuticals and personal care products in wastewater treatment. Environ. Sci. Technol. 38: 392A-399A.

United Nations (2011) United Nations, Department of Economic and Social Affairs, Population Division. World Contraceptive Use 2010 (POP/DB/CP/Rev2010). http://www.un.org/esa/population/publications/wcu2010/WCP_2010/Data.html

US EPA (2008a) Aquatic Life Criteria for Contaminants of Emerging Concern. Part I. General Challenges and Recommendations and Aquatic Life Criteria for Contaminants of Emerging Concern. Part II. Illustration Of Recommendations Using Data For 17α-Ethynylestradiol (Ee2). Draft document. http://water.epa.gov/scitech/swguidance/standards/upload/2008_06_03_criteria_sab-emergingconcerns.pdf

US EPA (2008b) ECOSAR 2008. http://epa.gov/oppt/newchems/tools/21ecosar.htm

US EPA (2010a) Treating Contaminants of Emerging Concern. A Literature Review Database. EPA-820-R-10-002. http://water.epa.gov/scitech/swguidance/ppcp/upload/cecliterature.pdf. Database: http://water.epa.gov/scitech/swguidance/ppcp/upload/cecremoval.mdb

US EPA (2010b) Endocrine Disruptor Screening Program (EDSP). Website publication http://www.epa.gov/endo/

US EPA (2011a) Pesticides Industry Sales and Usage 2006 and 2007. Market Estimates. http://www.epa.gov/opp00001/pestsales/07pestsales/market_estimates2007.pdf

US EPA (2011b) ECOTOX 2011. http://cfpub.epa.gov/ecotox

62

van Wezel, A.P., and Opperhuizen, A. (1995) Narcosis due to environmental pollutants in aquatic organisms: residue-based toxicity, mechanisms, and membrane burdens. Crit. Rev. Toxicol. 25: 255-279.

Vrkoslavova, J., Demnerova, K., Mackova, M., Zemanova, T., Macek, T., Hajslova, J., Pulkrabova, J., Hradkova, P., and Stiborova, H. (2010) Absorption and translocation of polybrominated diphenyl ethers (PBDEs) by plants from contaminated sewage sludge. Chemosphere 81: 381-386.

Wagner, M., and Oehlmann, J. (2009) Endocrine disruptors in bottled mineral water: total estrogenic burden and migration from plastic bottles. Environmental Science and Pollution Research International 16: 278-286.

Walker, C.H., Hopkin, S.P., Sibly, R.M., and Peakall, D.B. (2006) Principles of Ecotoxicology. 3rd ed. Taylor and Francis. Boca Raton, FL, USA.

Wittmer, I.K., Scheidegger, R., Bader, H.P., Singer, H., and Stamm, C. (2011) Loss rates of urban biocides can exceed those of agricultural pesticides. Sci. Total Environ. 409: 920-932.

Worldwatch Institute (2011) http://www.worldwatch.org/node/485

Wu, C., Spongberg, A.L., Witter, J.D., Fang, M., and Czajkowski, K.P. (2010) Personal Care Products by Soybean Plants from Soils Applied with Biosolids and Irrigated with Contaminated Water. Environ. Sci. Technol. 44: 6157-6161.

Yamamoto, H., Nakamura, Y., Kitani, C., Imari, T., Sekizawa, J., et al. (2007) Initial ecological risk assessment of eight selected human pharmaceuticals in Japan. Environ. Sci. 14: 177-93.

Ying, G.G., Williams, B., and Kookana, R. (2002) Environmental fate of alkylphenols and alkylphenol ethoxylates - a review. Environ. Int. 28: 215-226.

Yu, J.T., Bouwer, E.J., and Coelhan, M. (2006) Occurrence and biodegradability studies of selected pharmaceuticals and personal care products in sewage effluent. Agric. Water Manage. 86: 72-80.

Yu, J., Hu, J., Tanaka, S., and Fujii, S. (2009) Perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) in sewage treatment plants Water Res. 43: 2399-2408.