Pharmaceutical antibiotic compounds in soils - a review - Susane.info

J. Plant Nutr. Soil Sci. 2003, 166, 145±167 145 

Pharmaceutical antibiotic compounds in soils ± a review 

Sören Thiele-Bruhn 

Institute of Soil Science and Plant Nutrition, University of Rostock, D-18051 Rostock, Germany 

Accepted 11 February 2003 

Summary ± Zusammenfassung 

Antibiotics are highly effective, bioactive substances. As a result of their consumption, excretion, and persistence, they are disseminated mostly 

via excrements and enter the soils and other environmental compartments. Resulting residual concentrations in soils range from a few lg upto 

gkg ±1 and correspond to those found for pesticides. Numerous antibiotic molecules comprise of a non-polar core combined with polar 

functional moieties. Many antibiotics are amphiphilic or amphoteric and ionize. However, physicochemical properties vary widely among 

compounds from the various structural classes. Existing analytical methods for environmental samples often combine an extraction with acidic 

buffered solvents and the use of LC-MS for determination. In soils, adsorption of antibiotics to the organic and mineral exchange sites is mostly 

due to charge transfer and ion interactions and not to hydrophobic partitioning. Sorption is strongly influenced by the pH of the medium and 

governs the mobility and transport of the antibiotics. In particular for the strongly adsorbed antibiotics, fast leaching through soils by macropore 

or preferential transport facilitated by dissolved soil colloids seems to be the major transport process. Antibiotics of numerous classes are 

photodegraded. However, on soil surfaces this process if of minor influence. Compared to this, biotransformation yields a more effective 

degradation and inactivation of antibiotics. However, some metabolites still comprise of an antibiotic potency. Degradation of antibiotics is 

hampered by fixation to the soil matrix; persisting antibiotics were already determined in soils. Effects on soil organisms are very diverse, 

although all antibiotics are highly bioactive. The absence of effects might in parts be due to a lack of suitable test methods. However, dose and 

persistence time related effects especially on soil microorganisms are often observed that might cause shifts of the microbial community. 

Significant effects on soil fauna were only determined for anthelmintics. Due to the antibiotic effect, resistance in soil microorganisms can be 

provoked by antibiotics. Additionally, the administration of antibiotics mostly causes the formation of resistant microorganisms within the treated 

body. Hence, resistant microorganisms reach directly the soils with contaminated excrements. When pathogens are resistant or acquire 

resistance from commensal microorganisms via gene transfer, humans and animals are endangered to suffer from infections that cannot be 

treated with pharmacotherapy. The uptake into plants even of mobile antibiotics is small. However, effects on plant growth were determined for 

some species and antibiotics. 

Pharmazeutische Antibiotika in Böden ± ein Überblick 

Antibiotika sind hochgradig wirksame, bioaktive Substanzen. Infolge ihrer Anwendung, Ausscheidung und Persistenz werden sie meist über die 

Exkremente in Böden und andere Umweltkompartimente eingetragen. Die resultierenden Rückstandskonzentrationen in Böden im Bereich von 

wenigen lg bis zu g kg ±1 entsprechen in etwa denen von Pflanzenschutzmitteln. Die Molekülstruktur von Antibiotika besteht häufig aus einem 

unpolaren Kern und polaren Randgruppen. Viele Antibiotika sind amphiphil oder amphoter und bilden Ionen, jedoch weisen die zahlreichen 

Antibiotika unterschiedlicher Strukturklassen stark divergierende physikochemische Eigenschaften auf. In den vorliegenden Nachweismethoden 

für Umweltproben werden häufig sauer gepufferte Lösungsmittel zur Extraktion und eine Bestimmung mittels LC-MS kombiniert. Die 

Adsorption der Antibiotika an den organischen als auch an den mineralischen Bodenaustauschern erfolgt zumeist durch Ladungs- und 

Ionenwechselwirkungen und weniger durch hydrophobe Bindungen. Das Verteilungsverhalten hängt dabei entscheidend vom pH-Wert des 

Mediums ab und beeinflusst die Mobilität und Verlagerung der Antibiotika. Bei vielen, insbesondere stark adsorbierten Antibiotika sind v. a. 

schnelle Flieûvorgänge wie durch präferenziellen und Makroporenfluss sowie der Cotransport mit gelösten Bodenkolloiden von besonderer 

Bedeutung. Antibiotika vieler Strukturklassen können durch Licht abgebaut werden. Dieser Abbaupfad spielt auf Bodenoberflächen jedoch nur 

eine untergeordnete Rolle. Hingegen kommt es insbesondere durch biologische Transformationsprozesse zu einer intensiven Degradation und 

Inaktivierung der Antibiotika. Verschiedene Metaboliten weisen jedoch ebenfalls ein antibiotisches Potential auf. Der Abbau der Antibiotika wird 

durch die Festlegung in Böden gehemmt; dementsprechend wurde eine Persistenz verschiedener Antibiotika nachgewiesen. Trotz der starken 

bioaktiven Wirkung aller Antibiotika sind die festgestellten Effekte auf Bodenorganismen sehr unterschiedlich. Dies liegt nicht zuletzt an einem 

Mangel an geeigneten Testmethoden. In der Regel sind jedoch von Dosis und Wirkungsdauer abhängige Effekte insbesondere auf Mikroorganismen 

festzustellen, die zu Veränderungen der Mikroorganismenpopulation führen können. Lediglich durch Anthelmintika wurden deutliche Wirkungen auf 

Vertreter der Bodenfauna hervorgerufen. Infolge der antibiotischen Wirkung der Pharmazeutika kann eine Resistenzbildung bei Bodenorganismen 

ausgelöst werden. Zudem hat die Medikation von Antibiotika die Bildung resistenter Mikroorganismen bereits im behandelten Organismus zur Folge. 

DurchderenanschlieûendeAusscheidunggelangenresistenteKeimeauchdirektindieBöden.HandeltessichumresistentePathogeneoderkommtes 

zur Übertragung der Resistenzgene zwischen kommensalen und pathogenen Mikroorganismen, so besteht das erhebliche Risiko einer nicht 

therapierbaren Infektion von Mensch und Tier. Die Aufnahme selbst mobiler Antibiotika in die Pflanzen ist sehr gering. Dennoch wurden bei einigen 

Pflanzenarten Wirkungen von Antibiotika auf das Wachstum nachgewiesen. 

PNSS P02/01P 

1 Introduction 

Antibiosis is a natural chemical regulation mechanism among 

organisms, especially microorganisms. Consequently, biosynthesis 

of antibiotics also occurs in soils (Gottlieb, 1976). 

Today, a wide range of naturally occurring and of synthetic 

Correspondence: Dr. S. Thiele-Bruhn; E-mail: soeren.thiele@auf.unirostock.de 

antibiotics is frequently used for the therapy of infectious 

diseases in human and veterinary medicine (Gräfe, 1992). 

For this purpose, antibiotics are designed to act very 

effectively even at low doses and, in case of intra-corporal 

administration, to be completely excreted from the body after 

a short time of residence. Consequently, these substances 

are released to the environment. Therefore, residual 

concentrations of pharmaceutical antibiotics are found in 

ã 2003 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim 1436-8730/03/0204-145 $17.50+.50/0

146 Thiele-Bruhn J. Plant Nutr. Soil Sci. 2003, 166, 145±167 

the environment. Contamination of surface, ground and 

drinking water, of aquatic sediments and soils with 

pharmaceuticals have been reported (Richardson and Bowron, 

1985; Heberer and Stan, 1998; Hirsch et al., 1999; 

Kümmerer, 2001a; Hamscher et al., 2002a). Soil residues 

result mostly from the use of contaminated excrements as 

fertilizer on agricultural land. It has been estimated that loads 

of up to kilograms per hectare may enter agricultural soils and 

that a concentration level of antibiotics similar to pesticides is 

easily reached (van Gool, 1993; Winckler and Grafe, 2000). 

Due to surface runoff and leaching, soils can even act as a 

source of antibiotic contaminants for the aqueous environment 

(Alder et al., 2001). Residues of pharmaceutical 

antibiotics can provoke resistance in pathogens either directly 

or indirectly by transfer of plasmids from non-pathogens to 

pathogenic microorganisms (Wegener et al., 1998). The 

resulting antibiotic residues and resistant microorganisms 

can affect the natural soil microbial community and soil 

functions and may even harm animals and humans via the 

food chain (Richter et al., 1996; Kennedy et al., 2000). In 

addition, infections by resistant pathogens lower the 

efficiency of pharmacotherapies for humans and animals 

(Richter et al., 1996). 

However, no regulations exist for concentration limits of 

antibiotics in soils or soil water. Following the lead of the USA, 

an environmental risk assessment of veterinary pharmaceuticals 

was prescribed in the EU in 1998 with the EU directives 

81/852/EEC and 92/18/EEC (EMEA, 1997). Thereby, predicted 

environmental concentrations (PEC) are calculated 

with the help of a balancing model (Spaepen et al., 1997). 

The PEC are compared with predicted, biologically noneffective 

concentrations (PNEC). First, the exposition is 

evaluated and PEC are compared with trigger values. These 

trigger values have been set at 10 lg kg ±1 for the faeces of 

grazing livestock, at 100 lg kg ±1 for dung and soils and at 0.1 

lg l ±1 for groundwater. When PEC exceed these trigger 

values or antibiotics are directly applied to surface water for 

the treatment of fish, experimental testing in the tiered second 

phase becomes necessary. Legislation and methods for an 

environmental risk assessment of pharmaceuticals are dealt 

with in detail in the book edited by Kümmerer (2001b). There, 

alternative concepts to the model by Spaepen et al. (1997) 

are also presented. Therefore, this topic is not addressed in 

this review. 

The occurrence and effects of pharmaceuticals in different 

environmental compartments, especially water, were 

reviewed by Halling-Sùrensen et al. (1998), Daughton and 

Ternes (1999), and Kümmerer (2000). Another review deals 

with the adsorption of veterinary pharmaceuticals in soils 

(Tolls, 2001), while Richardson and Bowron (1985) and 

Hirsch et al. (1999) combined overviews on antibiotics in 

water with results from their own investigations. However, 

research and publications on these topics have increased 

remarkably in the last few years, justifying a review focusing 

on the input and fate of pharmaceutical antibiotics in soils. 

2 Consumption and physicochemical 

properties of antibiotics 

Antibiotics are defined as chemical compounds that are 

synthesized through the secondary metabolism of living 

organisms, with exceptions for semi- or completely synthetic 

substances. Antibiotics inhibit the activity of microorganisms, 

viruses, and eucaryotic cells, respectively (Lancini and 

Parenti, 1982). In human medicine, antibiotics pose the third 

biggest group among all pharmaceuticals making up more 

than 6 % of all prescriptions (Schwabe and Paffrath, 2001). In 

veterinary medicine, more than 70 % of all consumed 

pharmaceuticals are antibiotic agents (Halling-Sùrensen et 

al., 1998). In Europe, two thirds of all pharmaceutical 

Table 1: Annual consumption of antibiotics for veterinary medicine, especially livestock, in the European Union, European countries and 

regions of Germany (without coccidiostatics). 

Tabelle 1: Jährlicher Verbrauch an Antibiotika in der Veterinärmedizin, v.a. für landwirtschaftliche Nutztiere, in der Europäischen Union, 

europäischen Staaten und Regionen in Deutschland (ohne Kokzidiostatika). 

EU 

1999 a 

France 

1980 b 

Sweden 

1996 c 

Denmark 

1997 d 

Switzerland 

1997 e 

UK 

2000 f 

Weser-Ems 

1997 g 

Brandenburg 

1998/99 h 

t % t % t % t % t % t % t % t % t % 

Therapeutics 3902 625 20 57 14 437 88 j 6.6 j 10 j 

Mecklenburg- 

Vorpommern 

2001 i 

Tetracyclines 2575 66 117 19 2.7 13 13 23 1.0 7 228 52 40 57 4.6 69 6.4 41 

Sulfonamidesk 78 2 139 22 2.2 11 13 23 n.a. 94 22 14 21 0.9 14 2.5 16 

Aminoglycosides 1564 57 9 1.1 5 7.7 14 n.a. 12 3 7.1 10 0.2 3 0.2 2 

b-Lactams 351 9 50 8 n.a. 15 26 9.2 64 49 11 3.8 5 0.2 3 0.1 1 

Macrolides 468 12 37 6 1.5 7 1.7 3 0.3 2 41 9 0.2 0.3 0.01 0.1 0.1 1 

others 234 6 226 36 n.a. 6.5 11 3.8 27 12 3 4.5 6 0.7 10 0.8 8 

Ergotropics 786 n.a. n.a. 107 36 24 18 0.1 n.a. 

a FEDESA (2001), data for percentages of compound classes from 1997; b Espinasse (1993); c Mudd et al. (1998); d Halling-Sùrensen et al. 

(2002a); e Swiss Importers of Antibiotics (1998) cited in Alder et al. 2001; f NOAH (2002); g Winckler and Grafe (2000); h Linke and Kratz 

(2001); i Thiele-Bruhn et al. (2003a); j veterinary prescriptions for feed antibiotics only; k incl. trimethoprim; n.a. = no data available

J. Plant Nutr. Soil Sci. 2003, 166, 145±167 Antibiotic in soils 147 

antibiotics are used in human medicine and one third for 

veterinary purposes (FEDESA, 2001). Consequently, tons of 

antibiotic substances are consumed per year in industrialized 

countries (Tab. 1). Agricultural livestock, especially poultry 

and pigs, are treated with the majority of antibiotics while all 

other domestic animals receive only ca. 1 % of prescriptions 

within the EU (Ungemach, 2000). To date, the nontherapeutic 

use of antibiotics as growth promoters is almost 

completely restricted in the EU and consequently, the 

consumption of ergotropics is strongly declining. On the 

other hand, the annual application of therapeutic antibiotics in 

the EU increased from 1997 to 1999 by 12 %, i.e. to 3900 t 

(FEDESA, 2001). The antibiotic compound classes primarily 

administered in veterinary medicine are tetracyclines, 

sulfonamides, aminoglycosides, b-lactams, and macrolides 

(Tab. 1). In human medicine b-lactams, tetracyclines and 

macrolides are mostly prescribed (Schwabe and Paffrath, 

2001). In Germany, 250 different antibiotic and antimycotic 

agents are currently approved for use (Kümmerer, 2001b), a 

number which might be representative of other countries as 

well. 

Antibiotics define a multitude of heterogeneous compounds 

that are classified with regard to different fields of usage (e.g. 

antimycotics, antiinfectives, anthelmintics), in different struc- 

tural classes (e.g. nucleosides, tetracyclines) and exhibit 

different molecular structures and diverse chemical and 

physical properties (Gräfe, 1992). Most antibiotics tend to 

ionize depending on the pH of the medium; pK a values are 

associated with the different functional groups of the 

compounds. Ranges of physicochemical properties of 

important antibiotic compound classes are listed in Tab. 2. 

In the following section, their physicochemical properties are 

briefly described with information taken from Booth and 

McDonald (1988), Gräfe (1992), and Schadewinkel-Scherkl 

and Scherkl (1995), if not indicated otherwise. 

Tetracyclines 

Tetracyclines (TCs) are polyketides and comprise of a 

naphthacene ring structure. The TCs are amphoteric 

compounds as characterized by three pK a values. They are 

relatively stable in acids, but not in bases, and form salts in 

both media (Halling-Sùrensen et al., 2002b). The TCs form 

chelate complexes with divalent metal ions and b-diketones 

and strongly bind to proteins and silanolic groups (Oka et al., 

2000). Most TCs are sparingly water soluble, while the 

solubility of the corresponding hydrochlorides is much higher. 

Table 2: Representative pharmaceutical antibiotics and typical ranges of physicochemical properties from selected classes of antibiotics (Osol, 

1980; Gruber et al., 1990; Asuquo and Piddock, 1993; Nowara et al., 1997; Escribano et al., 1997; National Institutes of Health, 1999; Rabùlle 

and Spliid, 2000; Syracuse Research Corporation, 2001). 

Tabelle 2: Beispiele pharmazeutischer Antibiotika und typische Bereiche physikochemischer Eigenschaften von ausgewählten Strukturklassen 

der Antibiotika. 

Compound class Molar mass 

g mol ±1 

Water solubility 

mg l ±1 

log K ow pK a Henry¢s constant 

Pa l mol ±1 

Tetracyclines 444.5 ± 527.6 230 ± 52000 ±1.3 ± 0.05 3.3 / 7.7 / 9.3 1.7”10 ±23 ± 4.8”10 ±22 

chlortetracycline, oxytetracycline, tetracycline 

Sulfonamides 172.2 ± 300.3 7.5 ± 1500 ±0.1 ± 1.7 2 ± 3 / 4.5 ± 10.6 1.3”10 ±12 ± 1.8”10 ±8 

sulfanilamide, sulfadiazine, sulfadimidine, sulfadimethoxine, sulfapyridine, sulfamethoxazole 

Aminoglycosides 332.4 ± 615.6 10 ± 500 a ±8.1 ± ±0.8 6.9 ± 8.5 8.5”10 ±12 ± 4.1”10 ±8 

kanamycin, neomycin, streptomycin 

b-Lactams 334.4 ± 470.3 22 ± 10100 0.9 ± 2.9 2.7 2.5”10 ±19 ± 1.2”10 ±12 

penicillins: ampicillin, meropenem, penicillin G; cephalosporins: ceftiofur, cefotiam 

Macrolides 687.9 ± 916.1 0.45 ± 15 1.6 ± 3.1 7.7 ± 8.9 7.8”10 ±36 ± 2.0”10 ±26 

erythromycin, oleandomycin, tylosin 

Fluorquinolones 229.5 ± 417.6 3.2 ± 17790 ±1.0 ± 1.6 8.6 5.2”10 ±17 ± 3.2”10 ±8 

ciprofloxacin, enrofloxacin, flumequin, sarafloxacin, oxolinic acid 

Imidazoles 171.5 ± 315.3 6.3 ± 407 ±0.02 ± 3.9 2.4 2.3”10 ±13 ± 2.7”10 ±10 

fenbendazole, metronidazole, oxfendazole 

Polypeptides 499.6 ± 1038 not ± completely ±1.0 ± 3.2 negligible ± 2.8”10 ±23 

avermectin, bacitracin, ivermectin, virginiamycin 

Polyethers 670.9 ± 751.0 2.2”10 ±6 ± 3.1”10 ±3 5.4 ± 8.5 6.4 2.1”10 ±18 ± 1.5”10 ±18 

monensin, salinomycin 

Glycopeptides 

vancomycin 

1450.7 > 1000 not soluble in octanol 5.0 negligible 

Quinoxalinederivatives 

263.3 1.0”106 ±2.2 10 1.1”10 ±18 

olaquindox 

a gl ±1


The TCs strongly absorb light and thus, are susceptible to 

photodegradation (Mitscher, 1978). 

Sulfonamides 

Sulfonamides (SAs) are relatively insoluble in water. They are 

characterized by two pK a values indicating protonation of the 

amino group at a pH of 2 to 3 and deprotonation of the 

R 1 SO 2 NHR 2 moiety at a pH of 5 to 11 (Ingerslev and Halling- 

Sùrensen, 2000). In general, the amphoteric SAs behave as 

weak acids and form salts in strongly acidic or basic 

solutions. Mostly, SAs, substituted at the amino-N, have 

greatly reduced antibacterial activity. 

Aminoglycosides 

Antibiotics from the class of aminoglycosides are basic, 

strongly polar polycationic compounds. Their molecular 

structure is characterized by two or more amino sugars that 

are glycosidically bound to aminocyclitol. They are water 

soluble, mostly hydrophilic, and susceptible to photodegradation. 

b-Lactams 

Penicillins and cephalosporins are the two major sub-classes 

of the b-lactams. The antibiotic effect of penicillins is directly 

connected to the b-lactam ring. This ring is easily cleaved in 

acidic and basic media. Cephalosporins are derivatives of 7amino-cephalosporanic 

acid, condensed with a six-membered 

heterocycle in contrast to the five-membered heterocycle 

of penicillins. 

Macrolides 

Macrolides are defined as lactone structures with cycles of 

more than 10 C-atoms. Many macrolides are weak bases and 

are unstable in acids. Their water solubility varies considerably 

between the different derivatives. 

Fluorquinolones 

Most fluorquinolones (FQs), also known as quinolones, 

exhibit large chemical stability. They are insensitive to 

hydrolysis and increased temperatures, but are degraded 

by UV light. Their antibiotic potency depends mostly on the 

aromatic fluorine substituent at the C-6 position (Wetzstein, 

2001). 

3 Extraction and determination 

Numerous antibiotics are comprised of a non-polar core and 

polar functional groups (Juhel-Gaugain et al., 2000). They are 

sensitive to bases and strong acids and dissociate or 

protonate depending on the pH of the medium. Thereby, 

their distribution behavior changes considerably (Holten 

Lützhùft et al., 2000). Consequently, incomplete extraction 

of antibiotics with very polar and non-polar extractants and 

strong adsorption to polar and non-polar solid phase 

extractants (SPE) pose serious analytical problems. Thus, 

for the extraction of most antibiotics, the use of weakly acidic 

buffers in combination with organic solvents is recommended 

(Tab. 3). In food analysis, a 0.1 M EDTA-McIlvaine buffer (pH 

4.0) is often used (Weimann and Bojesen, 1999; Juhel- 

Gaugain et al., 2000). Cooper et al. (1998) and Kühne et al. 

(2000) suggested citric ethylacetate (pH 5.0) for the 

extraction of TCs. This method was adopted for soil samples 

by Hamscher et al. (2002a). At least in the case of TCs, this 

extractant yields better recovery rates from soil samples, 

although the solubility of numerous antibiotics of different 

structural classes is considerably smaller in pure ethylacetate 

as compared to methanol or DMSO (Salvatore and Katz, 

1993). Oka et al. (2000) reviewed techniques for the 

extraction and analysis of TCs and a compilation of methods 

can be found in Juhel-Gaugain et al. (2000). The use of 0.01 

M CaCl 2 to extract the mobile, non-adsorbed fraction of 

xenobiotics as prescribed by OECD (1997), cannot be 

recommended for TCs. These compounds form sparingly 

soluble complexes with Ca 2+ (Wessels et al., 1998). As an 

alternative to CaCl 2 , 0.1 M NH 4 NO 3 was successfully used 

(Thiele-Bruhn, unpublished data). 

In general, the sample clean-up is done by SPE or 0.45 lm 

filtration; extracts from centrifugation and liquid/liquid separation 

are usually concentrated by evaporation (Tab. 3). For 

SPE of SAs, reversed phases are very effective. In contrast, 

the usage of reversed phase materials for SPE of TCs 

requires pre-treatment of the solid phase with EDTAor 

silylating agents whereas for conditioning and elution, acidic 

buffered solvents are recommended (Oka et al., 1991; Zhu et 

al., 2001; Loke et al., 2002). The TCs bind so strongly to free 

silanol groups that they cannot be eluted by the usual organic 

solvents. Lindsey et al. (2001) suggested a macroporous 

copolymer combined with Na 2 EDTAas the chelating agent. 

To overcome the problems resulting from the amphoteric 

character of many antibiotics, new functionalized SPE 

materials that separate analytes by their hydrophobic as 

well as polar properties, e.g. hydrophilic-lipophilic balance 

cartridges (HLB) and mixed-mode HLB-cation exchange 

cartridges (MCX), appear to be advantageous (e.g. Kolpin 

et al., 2002). Golet et al. (2001) extracted FQs from 

wastewater with a mixed-mode silica based sorbent consisting 

of a non-polar phase and a strong cation exchanger 

interacting with the aromatic moiety of the core and the 

charged amino groups of the substituents. 

Antibiotics are usually separated by chromatographic techniques 

and subsequently detected. To avoid dissociation of 

the compounds or binding to free silanol groups from the 

widely used silica based chromatographic columns, dilute 

acids or weakly acidic buffers, e.g. phosphoric, citric, formic, 

and oxalic acid or EDTA, are commonly used (Oka et al., 

2000). Alternatively ion pair-chromatography is applied. 

Currently, the time and solvent consuming thin layer 

chromatography has been mostly replaced by high performance 

liquid chromatography combined with UV and Diode- 

Array Detection (HPLC-UV, -DAD). Increasingly, liquid 

chromatography with mass spectrometry (LC-MS) or tandem 

mass spectrometry (LC-MS/MS) (Thomashow et al., 1997; 

Oka et al., 2000; Hamscher et al., 2002a) with chemical


Table 3: Selected examples for extraction, separation and detection of antibiotic pharmaceuticals in environmental and food samples. 

Tabelle 3: Ausgewählte Beispiele zur Extraktion, Trennung und Bestimmung antibiotischer Pharmazeutika in Umweltproben und Lebensmitteln. 

further details a Reference 

Sample Extractant Clean-up Column type Eluent Detection 

& -limits 

Class/ 

compound 

Hirsch et al., 1998 

Campagnolo et 

al., 2002 

chloramphenicol: 

ESI-e MS/MS ESI+ d ; 


Table 3: Continued. 

Tabelle 3: Fortsetzung. 

further details a Reference 

Sample Extractant Clean-up Column type Eluent Detection 

& -limits 

Class/ 

compound 

Johannsen, 1991 

100 ll, post column 

derivat.: DMA/H SO 2 4 

UV 320 nm; 

10 lg kg ±1 

MeOH:phosphate 

buffer (pH 4)(9:1), 

0.7 ml min ±1 

Hypersil ODS 250”4; 

5 lm 

liquid/liquid MeCl 2 

animal feed MeOH/phosphate buffer 

(pH 4)(9:1) 

Polyethers 

(Salinomycin, 

Monensin, 

Lasalocid) 

Asukabe et al., 

1994 

pre column derivat.: 

1-(bromoacetyl) 

pyrene 

FLD ex 360 

em 420 nm 

MeOH:H O 97:3, 

2 

isocratic 

animal feed ACN SPE: Silica Develosil 5 C18 

250”4.6; 100 lm 

UV 270 nm Samuelsen et al., 

1994 

0.1 N NaOH centrifugation C18, 100”4.6; 3 lm 0.05 M H PO /ACN, 

3 4 

gradient 

marine 

sediment 

Sulfonamides 

UV 280 nm 40 C, 50 ll Ingerslev and 

Halling-Sùrensen, 

2000 

ACN:16.7 mM acetic 

acid/ (pH 5 with 4M 

NaOH) (17.5:82.5 ), 

1mlmin ±1 

Phenomenex ODS2 

C18 125”4.6; 5lm 

H O centrifugation, 

2 

0.45 lm filtration 

sewage 

sludge 

UV 265 nm 22 C, 10 ll Thiele, 2000 

0.01 M H PO / 3 4 

MeOH, 1 ml min ±1 , 

gradient 

soil MeOH SPE: C18 Nucleosil C18 250”4.6; 

100±5 lm 

25 C, 50 ll Haller et al., 2002 

H O:1mM NH OAc+10 % 

2 4 

ACN/ACN, 0.25 ml 

min ±1 UV; 

MS/MS ESI+ 

, gradient 

Nucleosil C18 125”3; 

100±5 lm 

pH 9 with KOH, EtOAc separation of EtOAc, 

0.45 lm filtration 

animal 

manure 

+Trimethoprim 

Lindsey et al., 

2001 

MS ESI+; 


ionization at atmospheric pressure (APCI) or electrospray 

ionization (ESI) is used (Oka et al., 1997; Carson et al., 

1998). Niessen (1998) reviewed the analysis of antibiotics by 

LC-MS and stated that this detection method is most 

sensitive for a multitude of compounds. Numerous applications 

already exist for LC-MS. Kamel et al. (1999) detected 

the highest sensitivity for TCs with LC-ESI-MS/MS in the 

positive ion mode and by addition of 1 % acetic acid in the 

mobile phase. In comparison, Lindsey et al. (2001) used 

ammonia formiate/formic acid in combination with a water/ 

methanol gradient. However, for FQs, fluorescence detection 

after derivatization and liquid chromatography was superior to 

LC-MS/MS (Golet et al., 2001). 

To date, the traditional, semi-quantitative detection of 

antibiotics by microbial inhibition tests such as agardiffusion 

and bioautography (Katz and Katz, 1983; Süûmuth et al., 

1987) is carried out only to supplement chemical analysis. 

However, for a complete assessment of antibiotics, not only 

their quantity, but also their antibiotic potential must be 

determined. For this purpose, new techniques that combine 

chemical extraction, chromatographic separation and determination 

by microbial assays have been developed (e.g. 

Sczesny et al., 2003). Hestbjerg Hansen et al. (2001) 

developed a biosensor for the selective detection of 

bioavailable and bioactive trace residues of TCs in soil. 

4 Input and concentrations in the soil 

environment 

Soils are a habitat and source of indigenous, antibiotics 

producing microorganisms (Gottlieb, 1976; Thomashow et 

al., 1997). Among numerous other soil microorganisms, 30 to 

50 % of actinomycetes isolated from soil are able to 

synthesize antibiotics (Topp, 1981). Such antibiotics, biosynthesized 

in situ, are found especially in the soil rhizosphere 

with concentrations of up to 5 lg g ±1 (Soulides, 1965; 

Lumsden et al., 1992; Shanahan et al., 1992). 

However, findings of pharmaceutical antibiotics in the 

environment increase. Like other pharmaceuticals, these 

compounds are optimized in their pharmacokinetics in such a 

way that they do not accumulate in the organism. After 

medication, they are mostly excreted as parent compounds, 

whereas metabolites might be also bioactive (Bouwman and 

Reus, 1994; Schadewinkel-Scherkl and Scherkl, 1995; 

Kümmerer et al., 2000). Excretion rates following the passage 

throughthegastro-intestinaltractareintherangeof40to90 %for 

SAs and TCs (Berger et al., 1986; Winckler and Grafe, 2001). 

Compilations of excretion rates were published by Zuccato et 

al. (2001), Halling-Sùrensen et al. (2001), and Jjemba (2002). 

Rates vary among the single antibiotic substances, the 

treated species and depend on the mode of application, as 

it was shown for SAadministered to pigs (Haller et al., 2002). 

4.1 Anthropogenic input 

Major portions of antibiotics are excreted after intra-corporal 

medication or are rinsed from the skin after dermal 

application. Consequently, antibiotics reach agricultural soils 

directly through grazing livestock or indirectly through the use 

of manure and sewage sludge as fertilizer (Jùrgensen and 

Halling-Sùrensen, 2000). In addition, wastewater and runoff 

from agricultural land are mainly responsible for the 

contamination of aquatic systems (Hirsch et al., 1999; Alder 

et al., 2001). Afurther significant source of antibiotics in the 

environment is their use in aquaculture for fish production. 

Here they are directly introduced into surface water (Römbke 

et al., 1996). Among other compounds, TCs, nitrofurans, and 

SAs are used mainly for this purpose (Löscher et al., 1994) 

and result in residual concentrations of several hundred mg 

kg ±1 in aquatic sediments (Jacobsen and Berglind, 1988; 

Samuelsen et al., 1992; Coyne et al., 1994). Flooding of 

shore soils with contaminated surface water may possibly 

yield an input of antibiotics. In contrast, environmental 

contamination due to the production and distribution of 

pharmaceuticals can mostly be excluded. However, severe 

ground water contamination following the deposition of 

pharmaceutical wastes from antibiotic production was 

reported by Holm et al. (1995). Since the 1950s, antibiotics 

have been used as pesticides, especially oxytetracycline and 

streptomycin, which are commonly used in fruit, vegetable, 

and ornamental plant production. In the USA, 0.5 % of the 

total antibiotic consumption of approximately 10,000 t is from 

the application to plants (McManus et al., 2002). In the vicinity 

of animal houses used for pig and poultry breeding, 

antibiotics were detected in dust from the exhaust air of the 

stable ventilation (Hamscher et al., 2002b; Thiele-Bruhn et 

al., 2003a). 

Intra-corporal degradation processes usually proceed in the 

faeces (Langhammer, 1989; Loke et al., 2000). In contrast, 

antibiotics not metabolized in the organism are often found as 

recalcitrant after excretion (Bouwman and Reus, 1994; 

Schadewinkel-Scherkl and Scherkl, 1995). Thus, several 

antibiotic compounds persist in the environment (Gavalchin 

and Katz, 1994; Kümmerer et al., 2000; Kühne et al., 2000) 

and are not transformed, e.g. by aeration of manure even at 

increased ambient temperature (Winckler and Grafe, 2001) 

or sewage water treatment (Richardson and Bowron, 1985). 

These substances are likely to reach aquatic sediments 

through waste water or agricultural soils after fertilization with 

manure and sewage sludge, respectively. 

4.2 Environmental concentrations 

The discussed sources of pharmaceuticals result in detectable 

residual concentrations in diverse environmental 

compartments and even in drinking water (Heberer and 

Stan, 1998; Hirsch et al., 1999). In the USA, a nationwide 

survey of pharmaceutical compounds revealed that among 

numerous other pharmaceuticals, a number of veterinary and 

human antibiotics were detected in 27 % of 139 river water 

samples at concentrations of up to 0.7 lg l ±1 (Kolpin et al., 

2002). In England, representative single substances from the 

classes of macrolides, SAs, and TCs were determined in river 

water in concentrations of ca. 1 lg l ±1 (Watts et al., 1982), a 

concentration that reduced aqueous microbial activity in 

biotests (Backhaus and Grimme, 1999). In marine sediment 

underneath fish farms, residual oxytetracycline concentrations 

of 500 to 4000 lg kg ±1 were commonly observed and


ca. 50 % of red rock crab (Cancer productus) collected after 

oxytetracycline application exceeded the US FDAlimit of 0.1 

lg g ±1 for seafood (Capone et al., 1996). 

The intra-corporal administration of antibiotics inevitably 

leads to residual concentrations in excrements. Consequently, 

antibiotics are frequently found in dung and manure. 

Manure samples from pigs contained up to 3.5 mg kg ±1 of 

SAs and up to 4 mg kg ±1 of TCs (Höper et al., 2002; 

Hamscher et al., 2002a; Sengelùv et al., 2003). Even larger 

concentrations of TCs were determined in dung from beef 

cattle and calves (Patten et al., 1980; Höper et al., 2002). 

Campagnolo et al. (2002) detected antimicrobial compounds 

from a multitude of different classes in all manure samples 

taken from eight pig farms, with the single substances often 

exceeding 100 lg l ±1 and the sum of all antibiotics 

approaching 1000 lg l ±1 . Additionally, surface and groundwater 

samples nearby the storage lagoons were often 

contaminated with antibiotics. Residual concentrations of 

antibiotics were estimated for agricultural soils, ranging for 

TCs from 450 to 900 lg kg ±1 (Winckler and Grafe, 2000), for 

macrolides from 13 to 67 lg kg ±1 and for FQs from 6 to 52 lg 

kg ±1 (Schüller, 1998). In soils under conventional landfarming 

fertilized with manure and monitored for two years, average 

concentrations of up to 199 lg kg ±1 tetracycline, 7 lg kg ±1 

chlortetracycline (Hamscher et al., 2002a), and 11 lg kg ±1 

sulfadimidine (Höper et al., 2002) were detected. Contamination 

of Swiss river water with veterinary antibiotics pointed to 

a runoff or leaching of antibiotics from soil to surface water 

(Alder et al., 2001). In contrast, Runsey et al. (1977) detected 

no antibiotics in weathered manure on pasture and soil and in 

runoff water. Additionally, b-lactams were rarely found in the 

environment, most probably due to the fast degradation of the 

chemically unstable lactam ring (Alder et al., 2001). 

5 Fate of antibiotics in soils 

5.1 Sorption and fixation 

The antibiotics of different structural classes vary considerably 

in their molecular structures and physicochemical 

properties (Tab. 2). Some substances are hydrophobic or 

non-polar, whereas others are completely water soluble or 

dissociate at pH values typical for soils. Thus, distribution 

coefficients (K d ) for the adsorption of antibiotics to soil 

material and aquatic sediments, e.g. as summarized by Tolls 

(2001), vary for SAs from 0.6 to 4.9, for TCs from 290 to 

1620, and for FQs from 310 to 6310 (Tab. 4). From the 

extractability, Katz and Katz (1983) deduced that the strength 

of sorption to soil increases in the following sequence: 

oxytetracycline = chlortetracycline £ bacitracin < tylosin < 

erythromycin < streptomycin. For most of the antibiotics, 

especially the stronger adsorbed TCs and FQs, desorption 

hysteresis is strong, and partition coefficients multiply in 

desorption experiments (Nowara et al., 1997; Rabùlle and 

Spliid, 2000). Yeager and Halley (1990) found that only 50 % 

of the adsorbed efrotomycin could be extracted, even with 

harsh solvents. However, sorption of antibiotics to soil 

minerals is weaker than to soil organic matter (SOM). 

Although adsorption of various SAs was stronger to the clay 

than to the sand size fraction of a soil, the opposite was true 

for the increase of the partition coefficients at the desorption 

step (Thiele et al., 2002). As shown for chlortetracycline, the 

adsorption of epimers and metabolites of an antibiotic may be 

considerably different from that of the parent compound and 

in the case of anhydro-chlortetracycline may be even lower 

(Tjùrnelund et al., 2000). This is of special importance when 

the more mobile metabolite still exhibits antimicrobial activity. 

Sorption of pharmaceutical antibiotics is especially influenced 

by soil pH (Holten Lützhùft et al., 2000), SOM (Langhammer, 

1989; Gruber et al., 1990), and soil minerals (Batchelder, 

1982). Pinck et al. (1961a) and Bewick (1979) reported a 

much stronger adsorption of aminoglycosides, TCs, and 

tylosin to expandable three layer clay minerals than to illite 

and kaolinite. Correspondingly, nearly the opposite sequence 

was found for the desorption of these antibiotics, while no 

release of aminoglycosides was determined (Pinck et al., 

1961b). Similar results were obtained for the adsorption of 

these pharmaceuticals and of streptomycin in soils characterized 

by the investigated different clay minerals (Pinck et 

al., 1961a; Soulides et al., 1962). It is assumed that besides 

adsorption, diffusion into porous soil particles also contributed 

to the fixation. An interlayer adsorption of antibiotics from 

various classes in the presence of montmorillonite was first 

described by Pinck et al. (1962). Correspondingly, the strong 

adsorption of FQs to soils, especially to clay minerals, was 

accompanied by an expansion of the spacing of montmorillonite 

(Nowara et al., 1997). The authors proposed 

coulombic interactions and the adsorption of anionic 

antibiotics via cation bridging to clay minerals as the main 

mechanism for FQ adsorption. Thereby, the deprotonated 

carboxylic group of FQ-carboxylic acids is fixed to the clay 

minerals while the sorption of the decarboxylated derivative is 

much smaller. 

The sorption and fixation of antibiotics is strongly governed by 

the property of numerous compounds to ionize depending on 

the pH of the medium (Yeager and Halley, 1990). Octanol/ 

water coefficients of ionizing compounds change considerably 

in a pH range around the acid dissociation constant 

(Holten Lützhùft et al., 2000). Electrostatic forces mostly 

drive the sorption of these derivatives to charged surfaces of 

mineral and organic exchange sites (Williams, 1982; Holten 

Lützhùft et al., 2000). The adsorption coefficients (K d )ofSAs 

increased from < 1 up to 30 when the soil pH decreased in the 

range of 8 to 4 (Boxall et al., 2002; Tolls et al., 2002). This 

was related to the ionization of the amphoteric sulfonamides. 

Correspondingly, the adsorption of TCs to humic substances 

and clay minerals was not only influenced by the pH, but also 

by the ionic strength of the medium (Sithole and Guy, 1987a, 

b). Oxytetracycline adsorption was 2.5 times greater when 

Ca 2+ was sorbed to clay minerals as compared to Na + 

(Sithole and Guy, 1987a), because TCs form reversible 

complexes with multivalent cations (Wessels et al., 1998). 

Chelate complexes of TCs preferentially involve the 

tautomeric C-11-C-12 b-diketone system that is formed under 

alkalinic conditions from the keto-enol molecule, while with 

decreasing pH, the dimethylamino group at C-4 position 

becomes increasingly involved (Loke et al., 2002). Sithole 

and Guy (1987a, b) proposed three major sorption mechan-


Table 4: Sorption coefficients of pharmaceutical antibiotics in soils, sediments and slurry. 

Tabelle 4: Adsorptionskoeffizienten pharmazeutischer Antibiotika in Böden, Sedimenten und Wirtschaftsdüngern. 

Antibiotic 

Class 

Antibiotic 

Compound 

Concentration 

±1 a lg g 

Sample 

soil: texture / pH / OC% 

K d 

lkg ±1 

K oc 

lkg ±1 

Kclay lkg ±1 

Reference 

Tetracyclines oxytetracycline sewage sludge / 6.5 / 37 b 3020 8160 a Holten-Lützhoft c 

33 ± 2000 mg g ±1 pig manure 6 h / 24 hd 83.2 / 77.6 195 Loke et al., 2002 

2.5 ± 50 loamy sand / 6.1 / 1.6 680 42500 Rabùlle and 

2.5 ± 50 sand / 5.6 / 1.4 670 47880 Spliid, 2000 

2.5 ± 50 sandy loam / 5.6 / 1.1 1026 93320 

2.5 ± 50 sand / 6.3 / 1.5 417 27790 

285 organic marine sediment 2590 Smith and 

10.9 organic marine sediment 663 Samuelsen, 1996 

Sulfonamides sulfachloropyridazine 0.05 ± 20 clay loam / 6.5 / 1.8 Boxall et al., 2002 

0.05 ± 20 sandy loam / 6.8 / 0.9 

sulfadimidine 0.2 ± 25 sand / 5.2 / 0.9 1.3e 139 Langhammer and 

0.2 ± 25 loamy sand / 5.6 / 2.3 3.5e 151 Büning-Pfaue, 

0.2 ± 25 sandy loam / 6.3 / 1.2 2.0e 170 1989 

0.2 ± 25 clay silt / 6.9 / 1.1 0.9e 80 

sulfamethazine 0.2 ± 25 sand / 5.2 / 0.9 1.2 174 Langhammer, 

0.2 ± 25 loamy sand / 5.6 / 2.3 3.1 125 1989 

0.2 ± 25 sandy loam / 6.3 / 1.2 2.0 208 

0.2 ± 25 clay silt / 6.9 / 1.1 1.0 82 

sulfapyridine 0.1 ± 500 silt loam / 7.0 / 1.6 1.6 101 Thiele, 2000 

0.1 ± 500 silt loam / 6.9 / 2.4 7.4 308 

1.0 ± 10 silt loam / 7.0 / 1.6 3.5 217 Thiele et al., 2002 

sulfadiazine 1.0 ± 10 silt loam / 7.0 / 1.6 2.0 124 

sulfadimidine 1.0 ± 10 silt loam / 7.0 / 1.6 2.4 149 

sulfanilamide 1.0 ± 10 silt loam / 7.0 / 1.6 1.7 106 

sulfadimethoxine 1.0 ± 10 silt loam / 7.0 / 1.6 2.3 143 

clay-loam / 6.2 / 3.1 10e 323 Tolls et al., 2002 

sulfachloropyridazine clay-loam / 6.2 / 3.1 4e 129 

sulfadiazine clay-loam / 6.2 / 3.1 2.5e 81 

sulfadimidine clay-loam / 6.2 / 3.1 3e 97 

sulfaisoxazole clay-loam / 6.2 / 3.1 1.5e 48 

sulfathiazole clay-loam / 6.2 / 3.1 3e 97 

Macrolides tylosin 50 mg g ±1 kaolinite 0.7 Bewick, 1979 

50 mg g ±1 illite 3.9 

500 mg g ±1 montmorillonite 0.7 

500 mg g ±1 bentonite 3.1 

100 ± 2000 mg g ±1 pig manure 6 h / 24 hd 45.7 / 240 110 Loke et al., 2002 

Macrolides tylosin 1.25 ± 25 loamy sand / 6.1 / 1.6 128 7990 Rabùlle and 

1.25 ± 25 sand / 5.6 / 1.4 10.8 771 Spliid, 2000 

1.25 ± 25 sandy loam / 5.6 / 1.1 62.3 5660 

1.25 ± 25 sand / 6.3 / 1.5 8.3 553 

Fluorquinolones ciprofloxacin 250 lg l ±1 sewage sludge / 6.5 / 37b 417 1127b Halling- 

Sùrensen, 2000 

2 ± 200 loamy sand / 5.3 / 0.70 427 61000 Nowara et al., 

enrofloxacin 2 ± 200 clay / 4.9 / 1.63 3037 186340 330 1997 

2 ± 200 loam / 5.3 / 0.73 5612 768740 2240




Antibiotic 

Class 

Antibiotic 

Compound 

Concentration 

±1 a lg g 

isms for TCs: complexation by divalent cations, ion 

exchange, and hydrogen bridging from acidic groups of 

humic acids to polar groups of the TCs. 

Adsorption of antibiotics to SOM is strong and depends on 

the quantity and the composition of SOM as well, as it was 

shown for sulfapyridine (Thiele, 2000). When a Chernozem 

was physically dispersed and separated into particle size 

fractions, the adsorption of SAs to the clay size fraction with 

stable organo-mineral complexes was about two times 

greater than to the sand size fraction, which was mostly 

characterized by particulate organic matter of plant origin 

(Thiele et al., 2002). The adsorption of sulfapyridine was 

significantly correlated with the concentration of lipids and 

lignin dimers in the SOM of the particle fractions. 

Correspondingly, the adsorption of oxytetracycline increased 

with increasing aromaticity of organic soil components (Suan 

and Dmitrenko, 1994a). As for soil minerals, the adsorption of 

oxytetracycline to humic acid varies significantly with pH 

Sample 


K d 

lkg ±1 

K oc 

lkg ±1 

Kclay lkg ±1 

Reference 

2 ± 200 loamy sand / 6.0 / 1.23 1230 99980 2276 

2 ± 200 loam / 7.5 / 1.58 260 16510 140 

2 ± 200 loamy sand / 5.3 / 0.70 496 70910 2480 

decarboxylated 

enrofloxacin 

2 ± 200 loamy sand / 5.3 / 0.70 7.7 1100 

ofloxacin 2 ± 200 loamy sand / 5.3 / 0.70 309 44140 

Imidazoles metronidazole 100 ± 2000 mg g ±1 pig manure 6 h / 24 hd n.d. n.d. Loke et al., 2002 

1.25 ± 25 loamy sand / 6.1 / 1.6 0.67 42 Rabùlle and 

1.25 ± 25 sand / 5.6 / 1.4 0.54 39 Spliid, 2000 

1.25 ± 25 sandy loam / 5.6 / 1.1 0.62 56 

1.25 ± 25 sand / 6.3 / 1.5 0.57 38 

fenbendazole 0.5 ± 100 silt loam / 7.0 / 1.6 0.91 57 Thiele and 

0.5 ± 100 silt loam / 6.9 / 2.4 0.84 35 Leinweber, 2000 

Polypeptides avermectin 0.006 ± 2.17 clay loam / 6.6 / 4.8 147 5300 Gruber et al., 

0.006 ± 2.17 sand / 7.5 / 0.1 17.4 30000 1990 

0.006 ± 2.17 silt loam / 7.5 / 2.1 80.2 6600 

Quinoxaline- olaquindox 100 ± 2000 mg g ±1 pig manure 6 h / 24 hd 20.4 / 9.77 50 Loke et al., 2002 

derivatives 1.25 ± 25 loamy sand / 6.1 / 1.6 1.67 104 Rabùlle and 

1.25 ± 25 sand / 5.6 / 1.4 1.21 86 Spliid, 2000 

1.25 ± 25 sandy loam / 5.6 / 1.1 1.27 116 

1.25 ± 25 sand / 6.3 / 1.5 0.69 46 

Lipoglycosides efrotomycin silt loam / 7.5 / 2.1 18 1460 Yeager and 

loam / 6.7 / 2.5 8.3 580 Halley, 1990 

1.0 ± 135 sandy loam / 7.5 / 1.1 51 8000 

1.0 ± 135 clay loam / 5.0 / 4.6 290 11000 

Diamino- trimethoprim 500 lg l 

pyrimidines 

±1 sewage sludge / 6.5 / 37b 76 205b Halling-Sùrensen 

et al., 2002b 

a if not indicated otherwise b Koc estimates for OC in dry matter from own data; c Holten-Lützhùft and Halling-Sùrensen cited in Stuer-Lauridsen 

et al. (2000); d adsorption time; e data derived from figures; n.d. = not detectable 

(Sithole and Guy, 1987b). The TCs bind to humic acids and 

proteins especially via anionic functional groups (Loke et al., 

2002). In mildly alkalinic manure (pH 7.8), tylosin Ais partly 

positively charged thus, the resulting sorption is likely due to 

ion binding to negative groups of the manure particles, but not 

to complexation with metal ions (Loke et al., 2002). Since K oc 

was sufficiently estimated from K ow , a major contribution of 

hydrophobic partitioning was concluded. However, the K oc 

concept is not valid for the majority of polar antibiotics (Thiele 

et al., 2002). The much stronger sorption of TCs to dissolved 

organic matter (DOM) than expected from K ow clearly 

stresses that sorption is not attributable to hydrophobic 

partitioning, but ionic interactions and hydrogen bonds (Tolls, 

2001). In general, the adsorption of antibiotics like FQs and 

SAs to faeces that are rich in organic matter is strong 

(Marengo et al., 1997; Thiele-Bruhn and Aust, 2003). 

However, distribution coefficients for oxytetracycline and 

tylosin are smaller in manure than in soils (Loke et al., 

2002). Accordingly, Boxall et al. (2002) determined decreas-


ing K d values for sulfachloropyridazine with an increasing 

proportion of manure in soil. This was not related to the 

mobilizing effect of DOM, but to the pH effect of the alkalinic 

manure. Similar effects of manure additions to soil were 

found for other SAs (Thiele-Bruhn and Aust, 2003). However, 

the slurry used in the latter investigation was acidic and 

above a manure to soil ratio of 1:10, K d values strongly 

increased. 

Adsorption of most antibiotics to soils is fast. Efrotomycin, a 

lipoglycoside, and SAs reach sorption equilibrium in soil after 

several hours (Langhammer and Büning-Pfaue, 1989; Yeager 

and Halley, 1990; Thiele, 2000). The adsorption of TCs to 

various exchange sites is characterized by two processes of 

different kinetics (Sithole and Guy, 1987b; Suan and 

Dmitrenko, 1994b) that can be interpreted as a fast initial 

adsorption to outer surfaces, followed by a penetration into 

interlayers of clay minerals and micropores. In contrast, for 

salinomycin surface adsorption to SOM is energetically 

preferred to incorporation into the humic substance as was 

revealed from molecular modelling by computational chemistry 

(Schulten, 2002). 

Adsorption mostly reduces the antibiotic potency of the 

compounds (Ingerslev and Halling-Sùrensen, 2000). It is 

assumed that this is especially the case when the bioactive 

functionality associates with the exchange sites (Wessels et 

al., 1998; Thiele, 2000). Correspondingly, desorption yields a 

reactivation of the antimicrobial potency (Samuelsen et al., 

1994; Halling-Sùrensen et al., 2002b). However, sorption or 

fixation does not necessarily result in a complete elimination 

of the antimicrobial activity (Halling-Sùrensen et al., 2003). 

5.2 Mobility and transport 

Field investigations revealed that point sources caused 

ground water contamination resulting from transport of 

antibiotics through soil as determined in the vicinity of manure 

lagoons (Campagnolo et al., 2002) and at a disposal site of a 

pharmaceutical plant (Holm et al., 1995). Adiffuse contamination 

of surface water by antibiotic leaching from 

agricultural soils has also been reported (Alder et al., 2001). 

In contrast, antibiotics were only detected in a small number 

of ground water samples from regions with intensive livestock 

production (Hirsch et al., 1999). However, to date, only a few 

systematic investigations related to the mobility and transport 

of antibiotics in soil exist. Numerous antibiotics have a low 

water solubility, they are relatively polar, and strongly retarded 

in soils (Tab. 2, Tab. 4). It is assumed that significant transport 

of such antibiotics like TCs is restricted to fast preferential 

and macropore flow or is facilitated by co-transport with 

mobile colloids like DOM. In submerged marine sediment, 

oxytetracycline transport within 220 days was restricted to 2 to 

4cm(Samuelsen et al., 1992). Also, leaching of avermectin in 

soil columns was small but increased significantly with 

preferential flow within the cracks of a structured silt loam 

(Gruber et al., 1990). In accordance with results from sorption 

experiments, the weakly adsorbing olaquindox completely 

leached through soil columns, while the stronger adsorbing 

tylosin was retained in different depths depending on the soil 

properties (Rabùlle and Spliid, 2000). No transport was found 

for oxytetracycline. Soil sorptive properties also governed the 

extent of sulfadimidine translocation in 30 cm soil columns 

filled with three soil materials of different properties 

(Langhammer and Büning-Pfaue, 1989). In contrast, sulfachloropyridazine 

was not leached through a sandy soil, while 

for a structured clay soil, rapid preferential transport of the SA 

into drainage water was observed within seven days (Boxall 

et al., 2002). 

5.3 Degradation and inactivation 

Numerous antibiotic compounds such as FQs, SAs, and TCs 

are susceptible to photodegradation (e.g. Burhenne et al., 

1997; Halling-Sùrensen et al., 2003). Accordingly, chlortetracycline 

was degraded on the surface of a Chernozem at a 

DT 50 of 5.8 days (Thiele-Bruhn et al., 2003b). However, under 

similar conditions no significant abiotic degradation of 

fenbendazole and sulfapyridine was determined. It is well 

known that photodecomposition in water already declines 

with increasing water depth and turbidity (Lunestad et al., 

1995). Thus, it was assumed that photodegradation has no 

significant effect on the concentration of antibiotics in soils, 

especially when they are spread onto soils as contaminants 

in sludge or slurry. As competing processes, fixation to and 

penetration into voids of the soil solids protect antibiotics from 

photodecomposition. Hydrolysis, another major abiotic process, 

also yields transformation of antibiotics (Halling- 

Sùrensen, 2000). The concentration of chlortetracycline 

aged in sterile soil possibly declined due to this process, 

while the extractable concentrations of fenbendazole and 

sulfapyridine did not change after an initial strong fixation 

(Thiele-Bruhn et al., 2003b). 

The degradation of xenobiotics in soils is mainly driven by 

microbial processes and numerous antibiotics are susceptible 

to enzymatic transformation reactions like oxidative decarboxylation 

and hydroxylation (Chen et al., 1997; McGrath et 

al., 1998; Al-Ahmad et al., 1999). Biodegradation was shown 

for ceftiofur-Na from the class of cephalosporins (Gilbertson 

et al., 1990). The antibiotic was quickly degraded in fortified 

cattle faeces, whereas no degradation was determined in 

sterile excrements. Thereby, intra-corporal metabolism 

proceeds in the environment. In mammals, antibiotics are 

mostly metabolized by a biphasic mechanism. First, functional 

groups are coupled to the molecule by monooxygenases, 

reductases, and hydrolases, followed by a covalent 

conjugation in the second phase, rendering the molecule 

more hydrophilic, excretable (Daughton and Ternes, 1999) 

and mostly antibiotic inactive (Halling-Sùrensen et al., 1998). 

The conjugation reactions are reversible though, and 

reactions back to the parent compound have been observed 

in the environment (Hussar et al., 1968; Langhammer, 1989; 

Halling-Sùrensen et al., 2002b). However, antibiotics usually 

further degrade in dung, manure, and soil (Halling-Sùrensen, 

2000; Ingerslev and Halling-Sùrensen, 2000;Wetzstein et al., 

2002).Additionsofmanureorsludge,containinghighnumbersof 

microorganisms, mostly result in increased biodegradation of 

antibiotics in soil (Ingerslev and Halling-Sùrensen, 2001; 

Ingerslev et al., 2001). Examples for degradation rates of 

antibiotics in manure and soils are listed in Tab. 5.


Degradation of antibiotics is governed by their molecular 

composition. Macrolides and penicillins are targets for fast 

degradation in soil (Gavalchin and Katz, 1994; Midtvedt, 

2001). The intense transformation of sarafloxacin (Marengo 

et al., 1997) and virginiamycin (Weerasinghe and Towner, 

1997) yielded numerous metabolites of minor concentration 

(< 10 %) and followed first order kinetics. In the presence of 

wood rotting fungi, FQs are inactivated through metabolism at 

the amine-position of the molecule while the heterocyclic ring 

is very persistent (Wetzstein, 2001). Thus, the xenobiotic 

fluorine moiety of FQs is quickly eliminated and the resulting 

metabolites exhibit no or strongly reduced antibiotic potential 

(Wetzstein et al., 1997). From the group of SAs, sulfadimethoxine 

is possibly demethylated in sediments (Samuelsen 

et al., 1994). 

However, fixation of antibiotic compounds to surfaces or in 

pores of the soil matrix may effectively protect them from 

biodegradation (Samuelsen et al., 1992; Gavalchin and Katz, 

1994) without alterations to their molecular structure (Halling- 

Sùrensen et al., 2002b). Consequently, bioavailable concentrations 

possibly decline below a threshold concentration 

for the initiation of biological degradation (Alexander, 1999). 

Thus, even compounds like TCs, which are reactive in 

standard solutions, persist in the soil for several months (van 

Gool, 1993; Höper et al., 2002). In aquatic sediments, no or 

only minor degradation of FQs, TCs, and SAs was 

determined over a period of 180 days (Samuelsen et al., 

1994; Hektoen et al., 1995). Besides of the physicochemical 

properties of the antibiotics, the extent and kinetics of 

degradation are also considerably affected by temperature 

(Gavalchin and Katz, 1994) and adsorption to soil (Weerasinghe 

and Towner, 1997) (Tab. 4). For the FQ enrofloxacin, 

it was shown that unspecific adsorption to SOM reduced 

degradation, although it was not completely stopped (Martens 

et al., 1996; Wetzstein et al., 1997). 

The products of biodegradation may still exhibit antimicrobial 

potential, as it was found for several metabolites of FQs 

(Weerasinghe and Towner, 1997; Marengo et al., 1997; 

Wetzstein et al., 2000). Furthermore, the decline in 

concentration of TCs and various antibiotics of other 

structural classes was not always mirrored by a decline in 

microbial toxicity (Halling-Sùrensen et al., 2002b, 2003). For 

example, tylosin Awas converted to the antibiotic tylosin B in 

acidic medium, whereas in neutral and alkaline medium, 

tylosin Aaldol was detected along with a number of polar 

decomposition products which were less bioactive (Halling- 

Sùrensen et al., 2003). 

Generally, the degradation of most xenobiotics is faster and 

more complete under aerobic as compared to anaerobic 

conditions. Correspondingly, degradation of oxytetracycline, 

tylosin, sulfadiazine, streptomycin, metronidazole, and olaquindox 

in activated sludge, soil and surface water was 

similar or slightly lower under anaerobic as compared to 

aerobic conditions, while ciprofloxacin was not degraded 

under anaerobic conditions (Ingerslev et al., 2001; Halling- 

Sùrensen et al., 2003; Tab. 5). Additionally, Ingerslev and 

Halling-Sùrensen (2000) demonstrated that acclimation of 

degrading organisms occurs in soil, when degradation times 

for SAs declined dramatically after respiking the soil with the 

same or a similar antibiotic. 

5.4 Effects on soil organisms and plants 

Pharmaceutical antibiotics are designed to affect mainly 

microorganisms. Hence, the toxic dose for microorganisms is 

often several magnitudes smaller than for higher organisms 

(Wollenberger et al., 2000). Accordingly, dose related effects 

on soil microorganisms were determined (Herron et al., 1998; 

Pfeiffer et al., 1998). Selected results are listed in Tab. 6. 

Effects and effective doses vary with time (Thiele and Beck, 

2001). When a Gleyic Podzol was incubated with tetracycline, 

the metabolic quotient was significantly affected by a 

concentration of 10 lg kg ±1 after 8 weeks (Höper et al., 

2002). After 16 weeks, effects were significant only at an 

initial concentration of 10 mg kg ±1 . Some antibiotics inhibit 

microorganisms (Colinas et al., 1994; Thiele and Beck, 

2001), while others promote their growth and activity (Patten 

et al., 1980; Hossain and Alexander, 1984). Antibiotics like 

streptomycin and cycloheximide are generally used to 

selectively inhibit growth of bacteria and fungi in soil 

experiments. Consequently, other pharmaceutical antibiotics 

cause changes in the composition of the indigenous soil 

microbial population as well (Ingham and Coleman, 1984; Mc 

Cracken and Foster, 1993). In this manner, even small 

extractable concentrations of oxytetracycline and sulfapyridine 

produced longer lasting significant effects (Thiele and 

Beck, 2001). 

In contrast, species of soil fauna were not affected by even 

excessive doses of antibiotics (Gomez et al., 1996; Herron et 

al., 1998; Baguer et al., 2000). While no effects of selected 

antibiotics on enchytreids and springtails were observed at 

environmentally relevant concentrations, the anthelmintic 

ivermectin significantly increased the mortality of springtails 

(Jensen et al., 2001b). Also, anthelmintics can change the 

fauna in and underneath cow-pats and hamper the dung 

decomposition (Madsen et al., 1988; McCracken and Foster, 

1993) and inhibit nematodes (Tomlinson et al., 1985) and 

earthworms (Gunn and Sadd, 1994; Tab. 6) 

The effects of antibiotics on organisms are essentially 

influenced by their bioavailability that depends on the soil 

properties, the availability of nutrients, and the presence of 

root exudates (da Gloria Britto de Oliveira et al., 1995; Herron 

et al., 1998). Multivalent cations inhibit the antibiotic potential 

of TCs and FQs (Froehner et al., 2000). First results 

confirming the reduction of antibiotic potency due to sorption 

and degradation in soils were published by Jefferys (1952). 

Degradation products of tylosin, sulfadiazine, streptomycin, 

ciprofloxacin, and olaquindox showed no significant potency 

in a soil bacterial assay (Halling-Sùrensen et al., 2003). 

However, a transformation of pharmaceuticals does not 

necessarily yield a decline in their antibiotic potential. Various 

metabolites of TCs still exhibited bacterial toxicity in sewage 

sludge and soil (Halling-Sùrensen et al., 2002b). 

The period over which antibiotics are effective depends on 

their persistence (Samuelsen et al., 1994). The antibiotic 

potential can increase with time (Dojmi di Delupis et al., 1992)


Table 5: Degradation of pharmaceutical antibiotics in soils and manure. 

Tabelle 5: Abbau pharmazeutischer Antibiotika in Böden und Wirtschaftsdüngern. 

Class Compound Concentration 

±1 a lg g 

Sample 


Degradation 

% 

Time 

d 

Reference 

Tetracyclines chlortetracycline cattle manure 24 84 Runsey et al., 1977 

5.6 sandy loam / 6.1+ manure 88 30 Gavalchin and Katz, 

1994 

4.7 lg kg ±1 soil 0 ca. 180 Hamscher et al., 

tetracycline 50±300 lg kg ±1 soil 0 ca. 180 2001 

10 poultry manure 65b 84 Jagnow, 1977 

10 manure + soil 100b 14 

pig manure, aerated 50 4.5 Kühne et al., 2000 

pig manure, non aerated 50 9 

20 ± 100 lg l ±1 pig manure 50 55±105 Winckler and Grafe, 

2001 

oxytetracycline soil + contam. manure 0 180 van Gool, 1993 

sediment slurry, aerobic 50 43.8 Ingerslev et al., 

2001 

Sulfonamides sulfanilamide various soils 0 14 Frankenberger and 

Tabatabai, 1982 

sulfabenzamide 250±1000 lg l ±1 manure / sludge 0 / 50c 28 / 0.4c Ingerslev and 

sulfadiazine ± º ± manure / sludge 0 / 50c 28 / 1.6c Halling-Sùrensen, 

sulfameter ± º ± manure / sludge 0 / 50c 28 / 0.4 2000 

sulfanilamide ± º ± manure / sludge 0 / 50c 28 / 3.8c sulfadimidine 1.0 loamy sand / 5.6 / 2.3 0.2 / 0.3d,e 64 Langhammer et al., 

1.0 clay silt / 6.9 / 1.1 0.3 / 0.7d,e 64 1990 

trimethoprim 500 lg l ±1 sewage sludge 50 22 ± 41 Halling-Sùrensen, 

2000 

Aminoglycosides streptomycin 5.6 sandy loam / 6.1+ manure 0 30 Gavalchin and Katz, 

b-Lactams penicillin 5.6 sandy loam / 6.1+ manure 0 30 1994 

ceftiotur clay loam 50e 22.2 Gilbertson et al., 

sand 50e 49.0 1990 

silty clay loam 50e 41.4 

mecillinam 500 lg l ±1 sewage sludge 50 0.5 ± 0.7 Halling-Sùrensen, 

2000 

Macrolides erythromecin 5.6 sandy loam / 6.1+ manure 25 30 Gavalchin and Katz, 

1994 

spiramycin poultry manure 70b 28 Jagnow, 1977 

tylosin 5.6 sandy loam / 6.1+ manure 0 30 Gavalchin and Katz, 

1994 

100 slurry + sand / 6.3 / 1.4 50 4.2 Ingerslev and 

Halling-Sùrensen, 

100 slurry+sandy loam /6.8/1.6 50 5.7 2001 

25 mg l ±1 pig manure, anaerobic 50 > 2.5 Loke et al., 2000 

Fluorquinolones sarafloxacin 3.4 loam / 7.9 0.58e 80 Marengo et al., 

3.4 sandy loam / 7.6 0.57e 80 1997 

3.4 silt loam / 7.6 0.49e 80 

enrofloxacin 10 sandy loam / 5.4 / 1.3 30.3e 56 Martens et al., 1996 

10 cattle manure + 

0.1-0.7 / 

basidiomycetes 

0.7±12.8e 56 Wetzstein et al., 

1997




Class Compound Concentration 

±1 a lg g 

from what may be due to bioaccumulation (Lo and Hayton, 

1981; Migliore et al., 1993). Bioaccumulation of 15 FQs was 

determined in E. coli, S. aureus and P. aeruginosa. However, 

the resulting intra-corporal concentrations were not correlated 

with the antibiotic effects of the pharmaceuticals (Asuquo and 

Piddock, 1993). 

In total, the knowledge about the ecotoxicity of antibiotics is 

still scarce, while their effects on humans and other mammals 

are well documented. There is a lack of special test methods 

and the adoption of unsuited methods produces erroneous 

results. No toxicity of antibiotics was observed with the 

bioluminescence test following the usual procedure, while 

after an extension of the duration of the assay, clear effects 

were determined (Backhaus and Grimme, 1999). Also, an 

Sample 


ciprofloxacin 10 mineral. media + 

wood rotting fungi 

Degradation 

% 

elongation of the substrate induced respiration test from 24 to 

48 hours was necessary to obtain dose response relationships 

from antibiotics in selected soil samples (Thiele and 

Beck, 2001). The nitrification inhibition test proved not to be 

suitable for antibiotic testing in wastewater (Alexy et al., 

2001). This might also explain for the nitrification rate in 

sewage sludge, manure, and soil, which was not affected by 

numerous antibiotics in various concentrations (Warman, 

1980; Patten et al., 1980; Gomez et al., 1996). 

5.4.1 Antibiotic resistance 

Time 

d 

Reference 

2.2 ± 35.3 e 56 Wetzstein et al., 

1999 

250 lg l ±1 sewage sludge 50 1.6 ± 2.5 Halling-Sùrensen, 

2000 

Imidazoles metronidazole 10 slurry + sand / 6.3 / 1.4 50 14.2 Ingerslev and 

10 slurry+sandy loam /6.8/1.6 50 15.0 Halling-Sùrensen, 

2001 

sediment slurry, aerobic 50 14 ± 75 Ingerslev et al., 

sediment slurry, anaerobic 50 74.5 2001 

Polypeptides bacitracin 5.6 sandy loam / 6.1+ manure 33 30 Gavalchin and Katz, 

1994 

Zn-bacitracin 25 poultry manure 90b 2 Jagnow, 1977 

25 manure + soil 100b 7 

virginiamycin 1.0 sandy silt / 8.2 50 87 Weerasinghe and 

1.0 silty sand / 6.3 50 116 Towner, 1997 

1.0 silty sand / 5.7 50 173 

1.0 silty clay loam / 6.1 21 64 

1.0 silty clay loam / 5.6 12 64 

1.0 clay loam / 5.4 18 64 

Polyethers monensin manure, anaerobic 30 ± 40 70 Donoho, 1984 

Phospholipoglycosides 

Quinoxalinederivatives 

flavomycin 5.6 sandy loam / 6.1+ manure 0 30 Gavalchin and Katz, 

1994 

flavophospholipol 10 poultry manure, aerobic 100b 119 Jagnow, 1977 

10 poultry manure, anaerobic 0b 119 

olaquindox 10 slurry + sand / 6.3 / 1.4 50 6.1 Ingerslev and 

10 slurry+sandy loam /6.8/1.6 50 5.8 Halling-Sùrensen, 

2001 

sediment slurry, aerobic 50 4.0 ± 7.0 Ingerslev et al., 

sediment slurry, anaerobic 50 21.5 2001 

Siderophores cyclosporin A200 ± 250 compost soil 50 60 Hübener et al., 1992 

a if not indicated otherwise; b concentration determined from antibiotic activity; c second values after respiking the antibiotic to the sample; 

d determined at 10 and 20 C, respectively; e14 CO2 determined from radio-labelled antibiotics 

Antibiotics released into the environment can provoke the 

formation of resistance, and even cross- and multiple 

resistance, in organisms (Nygaard et al., 1992; Wegener et


al., 1998; Al-Ahmad et al., 1999). Pathogens as well as 

commensal bacteria are affected, the latter constituting a 

potential reservoir of resistance genes for pathogenic 

bacteria. The transfer of such pathogens through the food 

chain is possible and consequently lowers the success of 

pharmacotherapies for curing humans and animals (Richter 

et al., 1996). The initiation of resistance is promoted by 

continuing a sublethal dosage of antibiotics (Gavalchin and 

Katz, 1994). This is put into effect by the repeated spreading 

of contaminated faeces onto agricultural soils (van Gool, 

Table 6: Effects of pharmaceutical antibiotics on soil organisms and plants. 

Tabelle 6: Wirkungen pharmazeutischer Antibiotika auf Bodenorganismen und Pflanzen. 

1993). The application of tetracycline contaminated manure 

induced antibiotic resistance in soil microorganisms that 

lasted for weeks (Fründ et al., 2000). Accordingly, tetracycline 

and oleandomycin resistant clostridia were significantly 

accumulated in manure, manure fertilized soils, and the 

ground water below (Huysman et al., 1993). However, 

survival of microorganisms in the presence of antibiotics is 

not necessarily due to acquired resistance. Many investigations 

revealed that numerous soil microorganisms have a 

natural tolerance towards antibiotics (Esiobu et al., 2002). 

Class Compound Habitat / Organism Effect / Inhibition Concentration 

±1 a lg g 

Reference 

Tetracyclines chlortetracycline Phaseolus vulgaris 70 % 10 mg l ±1 Batchelder, 1981 

oxytetracycline ± biomass 85 % 10 mg l ±1 

chlortetracycline ± biomass in sandy loam 51 % 160 

oxytetracycline ± biomass in sandy loam 56 % 160 

oxytetra. + penicillin hyphae of fungi: length and activity 48 % 10 Colinas et al., 1994 

oxytetracycline bacteria in sand soil 71 % 10 

chlortetracycline methane production in manure 100 % 18 mg l ±1 Fedler and Day, 

2002 

oxytetracycline silty sand / loamy sand ED SIR 10 0.81 / 0.93 Thiele and Beck, 

±ª± ED SIR 10 19.1 / 31.2 2001 

Sulfonamides sulfapyridine ± ª ± ED SIR 10 1.17 / 11.5 

±ª± ED SIR 10 0.05 / 6.20 

Tetracyclines tetracycline 13 microorg. strains MIC < 1 ± 1000 lg l ±1 van Dijck and van 

de Voorde, 1976 

8 soil microorg. strains MIC 10 lg l ±1 van Gool, 1993 

oxytetracycline springtails F. fimetaria LC /EC 10 10 >5000/>5000 Baguer et al., 2000 

earthworms A. caliginosa LC /EC 10 10 >5000 / 1954 

enchytreids E. crypticus LC /EC 10 10 >5000 / 3000 

Macrolides tylosin springtails F. fimetaria LC /EC 10 10 >5000 / 149 

earthworms A. caliginosa LC /EC 10 10 >5000 / 3306 

enchytreids E. crypticus LC /EC 10 10 2501 / 632 

Fluorquinolones ciprofloxacin sewage sludge bacteria EC50 0.61 mg l ±1 Halling-Sùrensen, 

Sulfonamides trimethoprim ± ª ± EC50 17.8 mg l ±1 2000 

mecillinam ± ª ± EC50 62.1 mg l ±1 

Aminoglycosides streptomycin ± ª ± EC50 0.47 mg l ±1 Halling-Sùrensen, 

Imidazoles metronidazole ± ª ± NOEC 100 mg l ±1 2001 

Macrolides tylosin ± ª ± EC50 54.7 mg l ±1 

Pleuromutilin tiamulin ± ª ± EC50 14.3 mg l ±1 

Fluorquinolones oxolinic acid ± ª ± EC50 0.10 mg l ±1 

Quinoxalines olaquindox ± ª ± EC50 95.7 mg l ±1 

û-Lactam penicillin G ± ª ± EC50 84.6 mg l ±1 

Sulfonamides sulfadiazine ± ª ± NOEC 60 mg l ±1 

Tetracyclines chlortetracycline ± ª ± EC50 0.4 mg l ±1 

oxytetracycline ± ª ± EC50 1.2 mg l ±1 

tetracycline ± ª ± EC50 2.2 mg l ±1




Class Compound Habitat / Organism Effect / Inhibition Concentration 

±1 a lg g 

Out of 36 strains of microorganisms from uncontaminated soil 

and water only seven were susceptible to 21 diverse 

antibiotics (van Dijck and van de Voorde, 1976). In particular, 

pseudomonads are often intrinsically resistant to antibiotics 

(Halling-Sùrensen et al., 2003). 

5.4.2 Input of resistant microorganisms into soils 

Resistance of soil microorganisms is not only provoked by the 

input of antibiotics into the environment. It appears to be even 

more important that resistant microorganisms are directly 

introduced with faeces into soils (Ali-Shtayeh et al., 1998). By 

dairy farm manure, resistance against ampicillin, penicillin, 

Reference 

Aminoglycosides streptomycin sewage sludge bacteria EC 0/10 h 50 0.42 / 0.61 mg l ±1 Halling-Sùrensen et 

Fluorquindones ciprofloxacin ± ª ± EC 0/10 h 50 0.025 / 0.008 mg l ±1 al., 2003 

Macrolides tylosin ± ª ± EC 0/10 h 50 17.5 / 24.9 mg l ±1 

Sulfonamides sulfadiazine ± ª ± EC 0/10 h 50 15.9 / 16.8 mg l ±1 

Tetracyclines oxytetracyclines ± ª ± EC 0/10 h 50 0.12 / 0.27 mg l ±1 

Aminoglycosides streptomycin soil fungi/bacteria/protozoa ±/0/+ b 3 Ingham et al., 1986 

± ª ± +/±/± b 30 

± ª ± no effect 1 Ingham and Colecycloheximid 

soil fungi/bacteria/protozoa no effect 1 man, 1984 

Anthelmintics ivermectin springtails LD50 10 Jensen et al., 2000 

springtails reproduction EC10 0.5 

soil respiration + b 11 Pfeiffer et al., 1998 

substrate induced respiration 0 b 11 

DMSO-reduction 0 b 11 

Polyether monensin soil respiration EC50 176 

substrate induced respiration + b 176 

DMSO-reduction + b 176 

various oxytetracycline, tylosin, tiamulin, metronidazole, olaquindox Jensen et al., 2001a 

enchytreids LD10 > 1000 

springtails LD10 > 1000 

springtails reproduction EC10 100 

Sulfonamindes sulfadimethoxine Panicum miliaceum roots plant 2071 Migliore et al., 1995 

Panicum miliaceum stems concentrations 110 

Pisum sativum roots from 178 

Pisum sativum stems bioaccumulation 60.2 

Zea mays roots 269 

Zea mays stems 12.5 

carrot root / stem / leaf ± / ± / ± b 1mM Migliore et al., 1996 

corn root / stem / leaf ± / ± / ± b 1mM 

millet root / stem / leaf 0 / 0 / 0 b 1mM 

pea root / stem / leaf ± / ± / ± b 1mM 

a b if not indicated otherwise; ± = inhibiton, 0 = no effect, + = promotion 

tetracycline, vancomycin, and streptomycin was significantly 

increased in a garden soil to a frequency of 70 % (Esiobu et 

al., 2002). The medication of pigs with chlortetracycline was 

followed by the excretion of microorganisms resistant 

towards chlortetracycline and a multitude of other antibiotics 

(Langlois et al., 1978). Ahigh prevalence of resistance 

against various antibiotics was determined in numerous 

faeces samples from pigs in the Netherlands and Sweden 

(van den Bogaard et al., 2000). Differences in the extent of 

resistance were related to the different intensity of antibiotic 

application in the two countries. Of special concern is the 

possible induction of antibiotic resistance in pathogens either 

directly or indirectly after transfer of genes encoding antibiotic 

resistance from non-pathogenic to pathogenic microorgan-


isms (Wegener et al., 1998). Transfer of resistance genes in 

soil was reviewed by Seveno et al. (2002). Mobile genetic 

elements conferring antibiotic resistance were readily 

obtained from microbial communities of environmental 

habitats (Smalla and Sobecky, 2002). As the major 

disseminating antibiotic resistance genes, bacterial IncQ 

plasmids were identified in pig manure (Smalla et al., 

2000). Additionally, genetic transfer in soil was shown to 

increase by fertilization with pig manure, since it consisted of 

plasmids with high mobility (Götz and Smalla, 1997). 

Accordingly, the extent of the increase of TC resistance in 

soil following manure fertilization was related to the amount of 

applied manure (Sengelùv et al., 2003). However, the 

resulting increased resistance level in soils and water was 

lower than in faeces (Langlois et al., 1978) and vanished 

within one month (Sengelùv et al., 2003). 

5.4.3 Uptake in and effects on plants 

The uptake and effects on plants varies considerably 

between reports and depends on the antibiotic substance 

and plant species (Patten et al., 1980; Langhammer, 1989; 

Migliore et al., 1995; Tab. 6). Yields and nutrient uptake by 

radish, wheat, and corn were increased in the presence of 

160 mg kg ±1 of TCs (Batchelder, 1982). In contrast, 

performance of pinto beans (Phaseolus vulgaris) was 

significantly reduced in a sandy loam, but not clay loam that 

is most likely due to the inhibition of root nodulation by 

rhizobia (Batchelder, 1982). The effects of antibiotics on 

plants were reviewed by Jjemba (2002). The author stated 

that negative impacts of therapeutic compounds on plants 

were mostly determined by in vitro experiments at concentrations 

that are unlikely to occur in field soils. Negative impacts 

of contaminated manure on field soils were most likely related 

to excessive nitrogen or heavy metals, but not antibiotics. 

Accordingly, the plant growth inhibiting effect of sulfadimethoxine 

was smaller in vivo than in vitro (Migliore et al., 

1996). This was related to a slow bioaccumulation of the 

antibiotic from the nutrient solution into the plant (Migliore et 

al., 1998) that may have been suppressed in soil by the aging 

of antibiotics. The translocation of 14 C-sulfadimidine from soil 

into maize plants declined from 15 to 3 % after 32 days of 

aging in soil (Langhammer et al., 1990). Additionally, the 

transfer from roots to shoots was less than 0.04 % of the total 

radiolabel. No uptake of TCs into pinto beans and coconut 

trees was observed even after direct application (McCoy, 

1976; Batchelder, 1981), although growth of beans in liquid 

cultures was significantly reduced at a concentration of 

10 mg l ±1 chlortetracycline and oxytetracycline (Batchelder, 

1981). However, endangerment of humans via the food chain 

is possible (Kennedy et al., 2000). 

6 Conclusions 

Residues of pharmaceutical antibiotics occur quiet often in 

soils. Hence, environmental risk assessment of pharmaceutical 

antibiotics is necessary and legally prescribed. However, 

there is still a considerable lack of knowledge, although the 

number of investigations on the input and fate of antibiotics in 

soils strongly increased in the last years. Especially little is 

known about the numerous antibiotics from classes less often 

administered to humans and animals. Only a few investigations 

exist on the mobility and resulting transport and 

bioavailability of antibiotics. Degradation pathways and 

kinetics and on the other hand the degree and causes for 

persistence of antibiotics in soils need further elucidation. 

Ecotoxicity resulting from long-term exposure to low doses 

and mixtures of compounds are a special problem associated 

with antibiotics in the environment. Additionally, the introduction 

of resistance as opposed to existing intrinsic resistance 

of soil microorganisms is not completely understood. Hence, 

a comprehensive evaluation of residual concentrations in the 

soil environment is not possible. Existing arbitrary trigger 

values are not scientifically based, while limit values are 

missing. Effective and harmonized analytical methods exist 

for food, but not for soils. Also, methods to evaluate 

ecotoxicity need to be adapted or developed. The actual 

used methods according to OECD guidelines were developed 

for the testing of pesticides and are often not suitable for the 

chemically different antibiotics. To reduce the risk of 

environmental contamination and to save the indispensable 

antibiotics as effective drugs against infectious diseases, a 

responsible use and reduction of the consumption is needed. 

Especially in agriculture, antibiotics must not be abused to 

compensate for insufficient hygiene in stables and for not 

species appropriate animal husbandry. 

Acknowledgments 

The help of S. L. Foran for language editing and of G. Jandl 

and M.-O. Aust for proof reading of the manuscript is 

gratefully acknowledged. 

Abbreviations 

Antibiotics: FQs fluoroquinolones; SAs sulfonamides; TCs 

tetracyclines 

Extraction: ACN acetonitrile; EtOAc ethylacetate; MeCl 2 

dichloromethane; MeOH methanol; NH 4 OAc ammonia 

oxalate; SPE solid phase extraction; C8/C18 reversed octyl/ 

octadecyl silica phases; HLB hydrophilic-lipophilic balance 

cartridges; MCX mixed-mode HLB-cation exchange cartridges 

Detection: HPLC high performance liquid chromatography; 

DAD diode array detector; FLD fluorescence light detector; 

UV ultraviolet light absorption detector; LC-MS liquid 

chromatography-mass spectrometry; MS/MS tandem mass 

spectrometer; APCI chemical ionisation at atmospheric 

pressure ; ESI electrospray ionization 

Soil: DOM dissolved organic matter; OC organic carbon; 

SOM soil organic matter 

Effects on organisms: EC/ED effective concentration/dose; 

EC 10 /EC 50 ± 10 %/50 % difference from control; LC/LD lethal 

concentration/dose; MIC minimal inhibiting concentration; 

NOEC no observable effect concentration


References 

Al-Ahmad, A., Daschner, F. D., and Kümmerer, K. (1999): 

Biodegradability of cefotiam, ciprofloxacin, meropenem, penicillin 

G, and sulfamethoxazole and inhibition of waste water bacteria. 

Arch. Environ. Contam. Toxicol. 37, 158±163. 

Alder, A. C., McArdell, C. S., Golet, E. M., Ibric, S., Molnar, E., 

Nipales, N. S., and Giger, W. (2001): Occurrence and fate of 

fluoroquinolone, macrolide, and sulfonamide antibiotics during 

wastewater treatment and in ambient waters in Switzerland. In C. 

G. Daughton and T. Jones-Lepp: Pharmaceuticals and Personal 

Care Products in the Environment: Scientific and Regulatory 

Issues, pp. 56±69. American Chemical Society, Washington, D.C. 

Alexander, M. (1999): Biodegradation and Bioremediation. Academic 

Press, San Diego, CA, p. 453. 

Alexy, R., Kümpel, T., Dörner, M., and Kümmerer, K. (2001): Effects 

of antibiotics against environmental bacteria studied with simple 

tests. SETAC Europe 11th Annual Meeting, 6±10 May, Madrid. 

Ali-Shtayeh, M. S., Jamous, R. M., and Abu-Ghdeib, S. I. (1998): 

Ecology of cycloheximide-resistant fungi in field soils receiving raw 

city wastewater or normal irrigation water. Mycopathologia 144, 

39±54. 

Asukabe, H., Murata, H., Harada, K., Suzuki, M., Oka, H., and Ikai, Y. 

(1994): Improvement of chemical-analysis of antibiotics. 21. 

Simultaneous determination of three polyether antibiotics in feeds 

using high-performance liquid-chromatography with fluorescence 

detection. J. Agric. Food Chem. 42, 112±117. 

Asuquo, A. E., and Piddock, L. J. (1993): Accumulation and killing 

kinetics of fifteen quinolones for Escherichia coli, Staphylococcus 

aureus and Pseudomonas aeruginosa. J. Antimicrob. Chemother. 

31, 865±880. 

Backhaus, T., and Grimme, L. H. (1999): The toxicity of antibiotic 

agents to the luminescent bacterium Vibrio fischeri. Chemosphere 

38, 3291±3301. 

Baguer, A. J., Jensen, J., and Krogh, P. H. (2000): Effects of the 

antibiotics oxytetracycline and tylosin on soil fauna. Chemosphere 

40, 751±757. 

Batchelder, A. R. (1981): Chlortetracycline and oxytetracycline 

effects on plant growth and development in liquid cultures. J. 

Environ. Qual. 10, 515±518. 

Batchelder, A. R. (1982): Chlortetracycline and oxytetracycline 

effects on plant growth and development in soil systems. J. 

Environ. Qual. 11, 675±678. 

Berger, K., Petersen, B., and Büning-Pfaue, H. (1986): Persistenz 

von Gülle-Arzneistoffen in der Nahrungskette. Arch. Lebensmittelhyg. 

37, 99±102. 

Bewick, M. W. M. (1979): The adsorption and release of tylosin by 

clays and soils. Plant Soil 51, 363±372. 

Booth, N. H., and McDonald, L. E. (1988): Veterinary Pharmacology 

and Therapeutics. Iowa State University Press, Ames, p. 1227. 

Bouwman, G. M., and Reus, J. A. W. A. (1994): Persistence of 

Medicines in Manure. Centre for Agriculture & Environment, CLM, 

163, p. 26. 

Boxall, A. B. A., Blackwell, P., Cavallo, R., Kay, P., and Tolls, J. 

(2002): The sorption and transport of a sulphonamide antibiotic in 

soil systems. Toxicol. Lett. 131, 19±28. 

Burhenne, J., Ludwig, M., Nikoloudis, P., and Spiteller, M. (1997): 

Photolytic degradation of fluoroquinolone carboxylic acids in 

aqueous solution. 1. Primary photoproducts and half-lives. 

Environ. Sci. Pollut. Res. 4, 10±15. 

Campagnolo, E. R., Johnson, K. R., Karpati, A., Rubin, C. S., Kolpin, 

D. W., Meyer, M. T., Esteban, J. E., Currier, R. W., Smith, K., Thug, 

K. M., and McGeehin, M. (2002): Antimicrobial residues in animal 

waste and water resources proximal to large-scale swine and 

poultry feeding operations. Sci. Total Environ. 299, 89±95. 

Capone, D. G., Weston, D. P., Miller, V., and Shoemaker, C. (1996): 

Antibacterial residues in marine sediments and invertebrates 

following chemotherapy in aquaculture. Aquaculture 145, 55±75. 

Carson, M. C., Ngoh, M. A., and Hadley, S. W. (1998): Confirmation 

of multiple tetracycline residues in milk and oxytetracycline in 

shrimp by liquid chromatography-particle beam mass spectrometry. 

J. Chromatogr. B 712, 113±128. 

Chen, Y., Rosazza, J. P., Reese, C. P., Chang, H. Y., Nowakowski, M. 

A., and Kiplinger, J. P. (1997): Microbial models of soil metabolism: 

biotransformations of danofloxacin. J. Ind. Microbiol. Biotechnol. 

19, 378±384. 

Colinas, C., Ingham, E., and Molina, R. (1994): Population responses 

of target and non-target forest soil-organisms to selected biocides. 

Soil Biol. Biochem. 26, 41±47. 

Cooper, A. D., Stubbings, G. W., Kelly, M., Tarbin, J. A., Farrington, 

W. H., and Shearer, G. (1998): Improved method for the on-line 

metal chelate affinity chromatography-high-performance liquid 

chromatographic determination of tetracycline antibiotics in animal 

products. J. Chromatogr. A812, 321±326. 

Coyne, R., Hiney, M., O¢Connor, B., Kerry, J., Cazabon, D., and 

Smith, P. (1994): Concentration and persistence of oxytetracycline 

in sediments under a marine salmon farm. Aquaculture 123, 

31±42. 

da Gloria Britto de Oliveira, R., Wolters, A. C., and van Elsas, J. D. 

(1995): Effects of antibiotics in soil on the population dynamics of 

transposon Tn5 carrying Pseudomonas fluorescens. Plant Soil 

175, 323±334. 

Daughton, C. G., and Ternes, T. A. (1999): Pharmaceuticals and 

personal care products in the environment: agents of subtle 

change? Environ. Health Perspect. 107, 907±938. 

Dojmi di Delupis, G., Macri, A., Civitareale, C., and Migliore, L. 

(1992): Antibiotics of zootechnical use: Effects of high and low 

dose contamination on Daphnia magna. Aquatic. Toxicol. 22, 

53±60. 

Donoho, A. L. (1984): Biochemical studies on the fate of monensin in 

animals and in the environment. J. Anim. Sci. 58, 1528±1539. 

EMEA (1997): Note for guidance: Environmental risk assessment for 

veterinary medicinal products other than GMO-containing and 

immunological products. Committee for Veterinary and Medicinal 

Products, London, p. 42. 

Escribano, E., Calpena, A. C., Garrigues, T. M., Freixas, J., 

Domenech, J., and Moreno, J. (1997): Structure-absorption 

relationships of a series of 6-fluoroquinolones. Antimicrob. Agents 

Chemother. 41, 1996±2000. 

Esiobu, N., Armenta, L., and Ike, J. (2002): Antibiotic resistance in 

soil and water environments. Int. J. Environ. Health Res. 12, 

133±144. 

Espinasse, J. (1993): Responsible use of antimicrobials in veterinary 

medicine: perspective in France. Vet. Microbiol. 35, 289±301. 

FEDESA (2001): Antibiotic use in farm animals does not threaten 

human health. Press release, Visby, Sweden. 13.06.2001. 

Fedler, C. B., and Day, D. L. (2002): Anaerobic digestion of swine 

manure containing an antibiotic inhibitor. Agric. Waste Utiliz. 

Manag. 523±530. 

Frankenberger, W. T., and Tabatabai, M. A. (1982): Transformations 

of amide nitrogen in soils. Soil Sci. Soc. Am. J. 46, 280±284. 

Froehner, K., Backhaus, T., and Grimme, L. H. (2000): Bioassays 

with Vibrio fischeri for the assessment of delayed toxicity. 

Chemosphere 40, 821±828.


Fründ, H.-C., Schlösser, A., and Westendarp, H. (2000): Effects of 

tetracycline on the soil microflora determined with microtiter plates 

and respiration measurement. Mitteilgn. Dtsch. Bodenkundl. 

Gesellsch. 93, 244±247. 

Gavalchin, J., and Katz, S. E. (1994): The persistence of fecal-borne 

antibiotics in soil. J. Assoc. Off. Anal. Chem. Int. 77, 481±485. 

Gilbertson, T. J., Hornish, R. E., Jaglan, P. S., Koshy, K. T., Nappier, 

J. L., Stahl, G. L., Cazers, A. R., Nappier, J. M., Kubicek, M. F., 

Hoffman, G. A., and Hamlow, P. J. (1990): Environmental fate of 

ceftiofur sodium, a cephalosporin antibiotic ± role of animal excreta 

in its decomposition. J. Agric. Food Chem. 38, 890±894. 

Golet, E. M., Alder, A. C., Hartmann, A., Ternes, T. A., and Giger, W. 

(2001): Trace determination of fluoroquinolone antibacterial agents 

in urban wastewater by solid-phase extraction and liquid 

chromatography with fluorescence detection. Anal. Chem. 73, 

3632±3638. 

Gomez, J., Mendez, R., and Lema, J. M. (1996): The effect of 

antibiotics on nitrification processes. Batch assays. Appl. Biochem. 

Biotechnol. 57±58, 869±876. 

Gottlieb, D. (1976): The production and role of antibiotics in soil. J. 

Antibiot. 29, 987±1000. 

Götz, A., and Smalla, K. (1997): Manure enhances plasmid 

mobilization and survival of Pseudomonas putida introduced into 

field soil. Appl. Environ. Microbiol. 63, 1980±1986. 

Gräfe, U. (1992): Biochemie der Antibiotika: Struktur ± Biosynthese ± 

Wirkmechanismus. Spektrum Akademischer Verlag, Heidelberg, p. 

389. 

Gruber, V. F., Halley, B. A., Hwang, S.-C., and Ku, C. C. (1990): 

Mobility of avermectin B in soil. J. Agric. Food Chem. 38, 

1a 

886±890. 

Gunn, A., and Sadd, J. W. (1994): The effect of ivermectin on the 

survival, behavior and cocoon production of the earthworm Eisenia 

fetida. Pedobiologia 38, 327±333. 

Haller, M. Y., Muller, S. R., McArdell, C. S., Alder, A. C., and Suter, M. 

J. F. (2002): Quantification of veterinary antibiotics (sulfonamides 

and trimethoprim) in animal manure by liquid chromatographymass 

spectrometry. J. Chromatogr. A952, 111±120. 

Halling-Sùrensen, B. (2000): Algal toxicity of antibacterial agents 

used in intensive farming. Chemosphere 40, 731±739. 

Halling-Sùrensen, B. (2001): Inhibition of aerobic growth and 

nitrification of bacteria in sewage sludge by antibacterial agents. 

Arch. Environ. Contam. Toxicol. 40, 451±460. 

Halling-Sùrensen, B., Nors Nielsen, S., Lanzky, P. F., Ingerslev, F., 

Holten Lützhùft, H. C., and Jùrgensen, S. E. (1998): Occurrence, 

fate and effects of pharmaceutical substances in the environment ± 

a review. Chemosphere 36, 357±393. 

Halling-Sùrensen, B., Jensen, J., Tjùrnelund, J., and Montforts, M. H. 

M. M. (2001): Worst-case estimations of predicted environmental 

soil concentrations (PEC) of selected veterinary antibiotics and 

residues used in Danish agriculture. In K. Kümmerer: Pharmaceuticals 

in the Environment: Sources, Fate, Effects and Risks. 

Springer, Berlin, pp. 143±157. 

Halling-Sùrensen, B., Nors Nielsen, S., and Jensen, J. (2002a): 

Environmental Assessment of Veterinary Medicinal Products in 

Denmark. Environmental Project No. 659 2002. Danish Environmental 

Protection Agency, Copenhagen, p. 97. 

Halling-Sùrensen, B., Sengelùv, G., and Tjùrnelund, J. (2002b): 

Toxicity of tetracyclines and tetracycline degradation products to 

environmentally relevant bacteria, including selected tetracyclineresistant 

bacteria. Arch. Environ. Contam. Toxicol. 42, 263±271. 

Halling-Sùrensen, B., Sengelùv, G., Ingerslev, F., and Jensen, L. B. 

(2003): Reduced antimicrobial potencies of oxytetracycline, tylosin, 

sulfadiazin, streptomycin, ciprofloxacin, and olaquindox due to 

environmental processes. Arch. Environ. Contam. Toxicol. 44, 

7±16. 

Hamscher, G., Sczesny, S., Höper, H., and Nau, H. (2001): 

Tierarzneimittel als persistente organische Kontaminanten in 

Böden. In Niedersächsisches Landesamt für Bodenforschung: 10 

Jahre Boden-Dauerbeobachtung in Niedersachsen. Hannover. 

Hamscher, G., Sczesny, S., Höper, H., and Nau, H. (2002a): 

Determination of persistent tetracycline residues in soil fertilized 

with liquid manure by high-performance liquid chromatography with 

electrospray ionization tandem mass spectrometry. Anal. Chem. 

74, 1509±1518. 

Hamscher, G., Pawelzick, H. T., Nau, H., and Hartung, J. (2002b): 

Detection of antibiotics in dust originating from a pig fattening farm. 

SETAC Europe 12th annual meeting. 12 ± 16 May 2002, Vienna. 

Heberer, T., and Stan, H.-J. (1998): Arzneimittelrückstände im 

aquatischen System. Wasser Boden 50, 20±25. 

Hektoen, H., Berge, J. A., Hormazabal, V., and Ynestad, M. (1995): 

Persistence of antibacterial agents in marine sediments. Aquaculture 

133, 175±184. 

Herron, P. R., Toth, I. K., Heilig, G. H. J., Akkermans, A. D. L., 

Karagouni, A., and Wellington, E. M. H. (1998): Selective effect of 

antibiotics on survival and gene transfer of streptomycetes in soil. 

Soil Biol. Biochem. 30, 673±677. 

Hestbjerg Hansen, L. H., Ferrari, B., Sùrensen, A. H., Veal, D., and 

Sùrensen, S. J. (2001): Detection of oxytetracycline production by 

Streptomyces rimosus in soil microcosms by combining whole-cell 

biosensors and flow cytometry. Appl. Environ. Microbiol. 67, 

239±244. 

Hirsch, R., Ternes, T., Haberer, K., and Kratz, K. L. (1999): 

Occurrence of antibiotics in the aquatic environment. Sci. Total 

Environ. 225, 109±118. 

Holm, J. V., Rugge, K., Bjerg, P. L., and Christensen, T. H. (1995): 

Occurrence and distribution of pharmaceutical organic-compounds 

in the groundwater downgradient of a landfill (Grindsted, 

Denmark). Environ. Sci. Technol. 29, 1415±1420. 

Holten Lützhùft, H. C., Vaes, W. H., Freidig, A. P., Halling-Sùrensen, 

B., and Hermens, J. L. (2000): 1-Octanol/water distribution 

coefficient of oxolinic acid: influence of pH and its relation to the 

interaction with dissolved organic carbon. Chemosphere 40, 

711±714. 

Höper, H., Kues, J., Nau, H., and Hamscher, G. (2002): Eintrag und 

Verbleib von Tierarzneimittelwirkstoffen in Böden. Bodenschutz 4, 

141±148. 

Hossain, A. K. M., and Alexander, M. (1984): Enhanzing soybean 

rhizosphere colonization by Rhizobium japonicum. Appl. Envir. 

Microbiol. 48, 468±472. 

Hübener, R., Dornberger, K., Zielke, R., and Gräfe, U. (1992): Abbau 

von Cyclosporin Adurch Bodenmikroorganismen. UWSF ± Z. 

Umweltchem. Ökotox. 4, 227±230. 

Hussar, D. A., Niebergall, P. J., Sugita, E. T., and Doluisio, J. T. 

(1968): Aspects of the epimerization of certain tetracycline 

derivatives. J. Pharm. Pharmacol. 20, 539±546. 

Huysman, E., van Renterghem, B., and Verstraete, W. (1993): 

Antibiotic resistant sulphite-reducing Clostridia in soil and groundwater 

as indicator of manuring practices. Water Air Soil Pollut. 69, 

243±255. 

Ingerslev, F., and Halling-Sùrensen, B. (2000): Biodegradability 

properties of sulfonamides in activated sludge. Environ. Toxicol. 

Chem. 19, 2467±2473. 

Ingerslev, F., and Halling-Sùrensen, B. (2001): Biodegradability of 

metronidazole, olaquindox, and tylosin and formation of tylosin


degradation products in aerobic soil-manure slurries. Ecotoxicol. 

Environ. Saf 48, 311±320. 

Ingerslev, F., Torang, L., Loke, M. L., Halling-Sùrensen, B., and 

Nyholm, N. (2001): Primary biodegradation of veterinary antibiotics 

in aerobic and anaerobic surface water simulation systems. 

Chemosphere 44, 865±872. 

Ingham, E. R., and Coleman, D. C. (1984): Effects of streptomycine, 

cycloheximide, fungizone, captan, carbofuran, cygon, and PCNB 

on soil microorganisms. Microb. Ecol. 10, 345±358. 

Ingham, E. R., Cambardella, C., and Coleman, D. C. (1986): 

Manipulation of bacteria, fungi and protozoa by biocides in 

lodgepole pine forest soil microcosms ± effects on organism 

interactions and nitrogen mineralization. Can. J. Soil Sci. 66, 

261±272. 

Jacobsen, P., and Berglind, L. (1988): Persistence of oxytetracycline 

in sediment from fish farms. Aquaculture 70, 365±370. 

Jagnow, G. (1977): Mikrobieller Abbau der Futtermittelantibiotika 

Zink-Bacitracin, Flavophospholipol, Spiramycin und von Tetracyclin 

in feucht gelagertem und in mit Boden vermischtem Hühnerkot. 

Landwirtsch. Forsch.1, 227±234. 

Jefferys, E. G. (1952): The stability of antibiotics in soils. J. Gen. 

Microbiol. 7, 295±312. 

Jensen, J., Krogh, P. H., and Sverdrup, L. (2001a): Environmental 

risk of veterinary medicines to soil fauna. SETAC Europe 11th Annual Meeting, 6±10 May, Madrid. 

Jensen, L. B., Baloda, S., Boye, M., and Aarestrup, F. M. (2001b): 

Antimicrobial resistance among Pseudomonas spp. and the 

Bacillus cereus group isolated from Danish agricultural soil. 

Environ. Int. 26, 581±587. 

Jjemba, P. K. (2002): The potential impact of veterinary and human 

therapeutic agents in manure and biosolids on plants grown on 

arable land: a review. Agric. Ecosyst. Environ. 93, 267±278. 

Johannsen, F. H. (1991): Zur Analytik von Wirkstoffen. 2. 

Bestimmung von Ionophor-Antibiotika in Futtermitteln und Lebensmitteln 

mittels Hochleistungsflüssigchromatographie und Nachsäulenderivatisierung. 

Agribiol. Res. 44, 79±89. 

Jùrgensen, S. E., and Halling-Sùrensen, B. (2000): Drugs in the 

environment. Chemosphere 40, 691±699. 

Juhel-Gaugain, M., McEvoy, J. D., and van Ginkel, L. A. (2000): 

Measurements for certification of chlortetracycline reference 

materials within the European Union: Standards, measurements 

and testing programme. Fres. J. Anal. Chem. 368, 656±663. 

Kamel, A. M., Brown, P. R., and Munson, B. (1999): Electrospray 

ionization mass spectrometry of tetracycline, oxytetracycline, 

chlorotetracycline, minocycline, and methacycline. Anal. Chem. 

71, 968±977. 

Katz, J. M., and Katz, S. E. (1983): Microbial assay systems for 

determining antibiotic residues in soils. J. Assoc. Anal. Chem. 66, 

640±645. 

Kennedy, D. G., Cannavan, A., and McCracken, R. J. (2000): 

Regulatory problems caused by contamination, a frequently 

overlooked cause of veterinary drug residues. J. Chromatogr. A 

882, 37±52. 

Kolpin, D. W., Furlong, E. T., Meyer, M. T., Thurman, E. M., Zaugg, S. 

D., Barber, L. B., and Buxton, H. T. (2002): Pharmaceuticals, 

hormones, and other organic wastewater contaminants in US 

streams, 1999±2000: Anational reconnaissance. Environ. Sci. 

Technol. 36, 1202±1211. 

Kühne, M., Ihnen, D., Möller, G., and Agthe, O. (2000): Stability of 

tetracycline in water and liquid manure. J. Vet. Med. A47, 

379±384. 

Kümmerer, K. (2000): Drugs, diagnostic agents and disinfectants in 

wastewater and water ± a review. Schriftenr. Ver. Wasser Boden 

Lufthyg. 105, 59±71. 

Kümmerer, K. (2001a): Drugs in the environment: emission of drugs, 

diagnostic aids and disinfectants into wastewater by hospitals in 

relation to other sources ± a review. Chemosphere 45, 957±969. 

Kümmerer, K. (2001b): Pharmaceuticals in the environment ± 

sources, fate, effects and risks. Springer, Berlin, p. 265. 

Kümmerer, K., Al Ahmad, A., and Mersch-Sundermann, V. (2000): 

Biodegradability of some antibiotics, elimination of the genotoxicity 

and affection of wastewater bacteria in a simple test. Chemosphere 

40, 701±710. 

Lancini, G., and Parenti, F. (1982): Antibiotics: An integrated View. 

Springer, New York, Heidelberg, p. 253. 

Langhammer, J.-P. (1989): Untersuchungen zum Verbleib antimikrobiell 

wirksamer Arzneistoffe als Rückstände in Gülle und im 

landwirtschaftlichen Umfeld. PhD thesis, Univ. Bonn, p. 138. 

Langhammer, J.-P., and Büning-Pfaue, H. (1989): Bewertung von 

Arzneistoff-Rückständen aus der Gülle im Boden. Lebensmittelchem. 

Gerichtl. Chem. 43, 103±113. 

Langhammer, J.-P., Führ, F., and Büning-Pfaue, H. (1990): Verbleib 

von Sulfonamid-Rückständen aus der Gülle in Boden und 

Nutzpflanze. Lebensmittelchemie 44, 93. 

Langlois, B. E., Cromwell, G. L., and Hays, V. W. (1978): Influence of 

chlortetracycline in swine feed on reproductive-performance and 

on incidence and persistence of antibiotic resistant enteric 

bacteria. J. Animal Sci. 46, 1369±1382. 

Lindsey, M. E., Meyer, T. M., and Thurman, E. M. (2001): Analysis of 

trace levels of sulfonamide and tetracycline antimicrobials in 

groundwater and surface water using solid-phase extraction and 

liquid chromatography/mass spectrometry. Anal. Chem. 73, 

4640±4646. 

Linke, I., and Kratz, W. (2001): Tierarzneimittel in der Umwelt ± 

Erhebung von Tierarzneimittelmengen im Land Brandenburg für 

den Zeitraum von Juli 1998 bis Juni 1999. LUABrandenburg, 

Potsdam, p. 34. 

Lo, I. H., and Hayton, W. L. (1981): Effects of pH on the accumulation 

of sulfonamides by fish. J. Pharmacokin. Biopharm. 9, 443±459. 

Loke, M. L., Ingerslev, F., Halling-Sùrensen, B., and Tjùrnelund, J. 

(2000): Stability of tylosin Ain manure containing test systems 

determined by high performance liquid chromatography. Chemosphere 

40, 759±765. 

Loke, M. L., Tjùrnelund, J., and Halling-Sùrensen, B. (2002): 

Determination of the distribution coefficient (log K ) of oxytetracy- 

d 

cline, tylosin A, olaquindox and metronidazole in manure. Chemosphere 

48, 351±361. 

Loke, M.-L., Jespersen, S., Vreeken, R., Halling-Sùrensen, B., and 

Tjùrnelund, J. (2003): Determination of oxytetracycline and its 

degradation products by high-performance liquid chromatographytandem 

mass spectrometry in manure containing anaerobic test 

systems. J. Chromatogr. B 783, 11±23. 

Löscher, W., Ungemach, F. R., and Kroker, R. (1994): Grundlagen 

der Pharmakotherapie bei Haus- und Nutztieren. P. Parey, Berlin, 

p. 437. 

Lumsden, R. D., Locke, J. C., Adkins, S. T., Walter, J. F., and Rideout, 

C. J. (1992): Isolation and localization of the antibiotic gliotoxin 

produced by Gliocladium virens from alginate prill in soil and 

soilless media. Phytopathol. 82, 230±235. 

Lunestad, B. T., Tore, B., Samuelsen, O. B., Fjelde, S., and Ervik, A. 

(1995): Photostability of eight antibacterial agents in seawater. 

Aquaculture 134, 217±225.


Madsen, E. L., Francis, A. J., and Bollag, J. M. (1988): Environmental 

factors affecting indole metabolism under anaerobic conditions. 

Appl. Environ. Microbiol. 54, 74±78. 

Marengo, J. R., Kok, R. A., OBrien, K., Velagaleti, R. R., and Stamm, 

J. M. (1997): Aerobic biodegradation of ( 14C)-sarafloxacin hydrochloride in soil. Environ. Toxicol. Chem. 16, 462±471. 

Martens, R., Wetzstein, H. G., Zadrazil, F., Capelari, M., Hoffmann, 

P., and Schmeer, N. (1996): Degradation of the fluoroquinolone 

enrofloxacin by wood-rotting fungi. Appl. Environ. Microbiol. 62, 

4206±4209. 

McCoy, R. E. (1976): Uptake, translocation, and persistence of 

oxytetracycline in coconut palm. Phytopathol. 66, 1038±1042. 

McCracken, D. I., and Foster, G. N. (1993): The effect of ivermectin 

on the invertebrate fauna associated with cow dung. Environ. 

Toxicol. Chem. 12, 73±84. 

McGrath, J. W., Hammerschmidt, F., and Quinn, J. P. (1998): 

Biodegradation of phosphonomycin by Rhizobium huakuii PMY1. 

Appl. Environ. Microbiol. 64, 356±358. 

McManus, P. S., Stockwell, V. O., Sundin, G. W., and Jones, A. L. 

(2002): Antibiotic use in plant agriculture. Annu. Rev. Phytopathol. 

40, 443±465. 

Midtvedt, T. (2001): Antibiotics in the Environment: Zinc bacitracin ± 

environmental toxicity and breakdown. In K. Kümmerer: Pharmaceuticals 

in the Environment: Sources, Fate, Effects and Risks. 

Springer, Berlin, pp. 77±79. 

Migliore, L., Brambilla, G., Grassitellis, A., and Dojmi di Delupis, G. 

(1993): Toxicity and bioaccumulation of sulphadimethoxine in 

Artemia (Crustacea, Anostraca). Int. J. Salt Lake Res. 2, 141±152. 

Migliore, L., Brambilla, G., Cozzolino, S., and Gaudio, L. (1995): 

Effect on plants of sulphadimethoxine used in intensive farming 

(Panicum miliaceum, Pisum sativum and Zea mays). Agric. 

Ecosyst. Environ. 52, 103±110. 

Migliore, L., Brambilla, G., Casoria, P., Civitareale, C., Cozzolino, S., 

and Gaudio, L. (1996): Effects of antimicrobials for agriculture as 

environmental pollutants. Fres. Envir. Bull. 5, 735±739. 

Migliore, L., Civitareale, C., Cozzolino, S., Casoria, P., Brambilla, G., 

and Gaudio, L. (1998): Laboratory models to evaluate phytotoxicity 

of sulphadimethoxine on terrestrial plants. Chemosphere 37, 

2957±2961. 

Mitscher, L. A. (1978): The Chemistry of the Tetracycline Antibiotics. 

Marcel Dekker, Basel, p. 330. 

Mudd, A. J., Lawrence, K., and Walton, J. (1998): Study of Sweden¢s 

model on antimicrobial use shows usage has increased since 1986 

ban. Feedstuffs 10. 

National Institutes of Health (1999): U.S. National Library of 

Medicine. Toxnet ± Toxicology Data Network. http://www.toxnet.nlm.nih.gov. 

Niessen, W. M. A. (1998): Analysis of antibiotics by liquid 

chromatography mass spectrometry. J. Chromatogr. A812, 53±75. 

NOAH (2002): Sales of antimicrobial Products used as veterinary 

Medicines, Growth Promoters and Coccidiostats in the UK in 2000. 

National Office of Animal Health, UK. 

Nowara, A., Burhenne, J., and Spiteller, M. (1997): Binding of 

fluoroquinolone carboxylic acid derivatives to clay minerals. J. 

Agric. Food Chem. 45, 1459±1463. 

Nygaard, K., Lunestad, B. T., Hektoen, H., Berge, J. A., and 

Hormazabal, V. (1992): Resistance to oxytetracycline, oxolinic 

acid and furazolidone in bacteria from marine sediments. 

Aquaculture 104, 31±36. 

OECD (1997): Adsorption/desorption using a batch equilibrium 

method. OECD guidelines for testing of chemicals, test guideline 

106, revised Draft Document. Organization for Economic Cooperation 

and Development, Paris. 

Oka, H., Ikai, Y., Kawamura, N., and Hayakawa, J. (1991): Limited 

survey of residual tetracyclines in tissues collected from diseased 

animals in Aichi prefecture, Japan. J. Assoc. Off. Anal. Chem. Int. 

74, 894±896. 

Oka, H., Ikai, Y., Ito, Y., Hayakawa, J., Harada, K., Suzuki, M., Odani, 

H., and Maeda, K. (1997): Improvement of chemical analysis of 

antibiotics. XXIII. Identification of residual tetracyclines in bovine 

tissues by electrospray high-performance liquid chromatographytandem 

mass spectrometry. J. Chromatogr. B 693, 337±344. 

Oka, H., Ito, Y., and Matsumoto, H. (2000): Chromatographic analysis 

of tetracycline antibiotics in foods. J. Chromatogr. A882, 109±133. 

Osol, A. (1980): Remington¢s Pharmaceutical Sciences. Mack 

Publishing Co., Easton, PA, p. 1928. 

Patten, D. K., Wolf, D. C., Kunkle, W. E., and Douglass, L. W. (1980): 

Effect of antibiotics in beef cattle feces on nitrogen and carbon 

mineralization in soil and on plant growth and composition. J. 

Environ. Qual. 9, 167±172. 

Pfeiffer, C., Emmerling, C., Schröder, D., and Niemeyer, J. (1998): 

Antibiotika (Ivermectin, Monensin) und endokrine Umweltchemikalien 

(Nonylphenol, Ethinylöstradiol) im Boden: Mögliche Auswirkungen 

von synthetischen Umweltchemikalien auf mikrobielle 

Eigenschaften eines landwirtschaftlich genutzten Bodens. 

Umweltwiss. Schadst. Forsch. 10, 147±153. 

Pinck, L. A., Holton, W. F., and Allison, F. E. (1961a): Antibiotics in 

soils: 1. Physico-chemical studies of antibiotic-clay complexes. Soil 

Sci. 91, 22±28. 

Pinck, L. A., Soulides, D. A., and Allison, F. E. (1961b): Antibiotics in 

soils: II. Extent and mechanisms of release. Soil Sci. 91, 94±99. 

Pinck, L. A., Soulides, D. A., and Allison, F. E. (1962): Antibiotics in 

soils: 4. Polypeptides and macrolides. Soil Sci. 94, 129±131. 

Rabùlle, M,. and Spliid, N. H. (2000): Sorption and mobility of 

metronidazole, olaquindox, oxytetracycline and tylosin in soil. 


Richardson, M. L., and Bowron, J. M. (1985): The fate of 

pharmaceutical chemicals in the aquatic environment. J. Pharm. 

Pharmacol. 37, 1±12. 

Richter, A., Löscher, W., and Witte, W. (1996): Leistungsförderer mit 

antibakterieller Wirkung: Probleme aus pharmakologisch-toxikologischer 

und mikrobiologischer Sicht. Prakt.Tierarzt 7, 603±624. 

Römbke, J., Knacker, T., and Stahlschmidt-Allner, P. (1996): 

Umweltprobleme durch Arzneimittel ± Literaturstudie. UBA Texte, 

Berlin, p. 341. 

Runsey, T. S., Miller, R. W., and Dinius, D. A. (1977): Residue content 

of beef feedlot manure after feeding diethylstilbestrol, chlortetracycline 

and Ronnel and the use of stirofos to reduce population of fly 

larvae in feedlot manure. Arch. Environ. Contam. Toxicol. 6, 

203±212. 

Salvatore, M. J., and Katz, S. E. (1993): Solubility of antibiotics used 

in animal feeds in selected solvents. J. Assoc. Off. Anal. Chem. Int. 

76, 952±956. 

Samuelsen, O. B. (1994): Simultaneous determination of 

ormethoprim and sulphadimethoxine in plasma and muscle of 

Atlantic salmon (Salmo salar). J. Chromatogr. B 660, 412±417. 

Samuelsen, O. B., Torsvik, V, and Ervik, A. (1992): Long-range 

changes in oxytetracycline concentration and bacterial resistance 

toward oxytetracycline in a fish farm sediment after medication. 

Sci. Total Environ. 114, 25±36. 

Samuelsen, O. B., Lunestad, B. T., Ervik, A., and Fjelde, S. (1994): 

Stability of antibacterial agents in an artificial marine aquaculture


sediment studied under laboratory conditions. Aquaculture 126, 

283±290. 

Schadewinkel-Scherkl, A.-M., and Scherkl, R. (1995): Antibiotika und 

Chemotherapeutika in der tierärztlichen Praxis. Gustav Fischer, 

Jena, Stuttgart, p. 153. 

Schüller, S. (1998): Anwendung antibiotisch wirksamer Substanzen 

beim Tier und Beurteilung der Umweltsicherheit entsprechender 

Produkte. 3. Statuskolloquium ökotoxikologischer Forschungen in 

der Euregio Bodensee, 3 ± 4 December 1998. 

Schulten, H. R. (2002): New approaches to the molecular structure 

and properties of soil organic matter: humic-, xenobiotic-, 

biological-, and mineral-bonds. In Violante, A., Huang, P. M., 

Bollag, J.-M., and Gianfreda, L.: Developments in Soil Science. 

Soil Mineral-Organic Matter-Microorganisms Interactions and 

Ecosystem Health 28A, Elsevier, Amsterdam, pp. 351±381. 

Schwabe, U., and Paffrath, D. (2001): Arzneiverordnungs-Report 

2001. Springer Verlag, Berlin, Heidelberg, p. 982. 

Sczesny, S., Nau, H., and Hamscher, G. (2003): Residue analysis of 

tetracyclines and their metabolites in eggs and in the environment 

by HPLC coupled with a microbiological assay and tandem mass 

spectrometry. J. Agric. Food Chem. 51, 697±703. 

Sengelùv, G., Agerso, Y., Halling-Sùrensen, B., Baloda, S. B., 

Andersen, J. S., and Jensen, L. B. (2003): Bacterial antibiotic 

resistance levels in Danish farmland as a result of treatment with 

pig manure slurry. Environ. Int. 28, 587±595. 

Seveno, N. A., Kallifidas, D., Smalla, K., van Elsas, J. D., Collard, J. 

M., Karagouni, A. D., and Wellington, E. M. H. (2002): Occurrence 

and reservoirs of antibiotic resistance genes in the environment. 

Rev. Med. Microbiol. 13, 15±27. 

Shanahan, P., Borro, A., O¢Gara, F., and Glennon, J. D. (1992): 

Isolation, trace enrichment and liquid chromatographic analysis of 

diacetylphloroglucinol in culture and soil samples using UV and 

amperometric detection. J. Chromatogr. A606, 171±177. 

Sithole, B. B., and Guy, R. D. (1987a): Models for oxytetracycline in 

aquatic environments. I. Interaction with bentonite clay systems. 

Water Air Soil Pollut. 32, 303±314. 

Sithole, B. B., and Guy, R. D. (1987b): Models for oxytetracycline in 

aquatic environments. II. Interactions with humic substances. 

Water Air Soil Pollut. 32, 315±321. 

Smalla, K., and Sobecky, P. A. (2002): The prevalence and diversity 

of mobile genetic elements in bacterial communities of different 

environmental habitats: insights gained from different methodological 

approaches. FEMS Microbiol. Ecol. 42, 165±175. 

Smalla, K., Heuer, H., Gotz, A., Niemeyer, D., Krogerrecklenfort, E., 

and Tietze, E. (2000): Exogenous isolation of antibiotic resistance 

plasmids from piggery manure slurries reveals a high prevalence 

and diversity of IncQ-like plasmids. Appl. Environ. Microbiol. 66, 

4854±4862. 

Smith, P., and Samuelsen, O. B. (1996): Estimates of the significance 

of out-washing of oxytetracycline from sediments under Atlantic 

salmon sea-cages. Aquaculture 144, 17±26. 

Soulides, D. A. (1965): Antibiotics in soils. VII. Production of 

streptomycin and tetracyclines in soils. Soil Sci. 100, 200±206. 

Soulides, D. A., Pinck, L. A., and Allison, F. E. (1962): Antibiotics in 

soils: V. Stability and release of soil-adsorbed antibiotics. Soil Sci. 

4, 239±244. 

Spaepen, K. R. I., van Leemput, L. J. J., Wislocki, P. G., and 

Verschueren, C. (1997): Auniform procedure to estimate the 

predicted environmental concentration of the residues of veterinary 

medicines in soil. Environ. Toxicol. Chem. 16, 1977±1982. 

Stuer-Lauridsen, F., Birkved, M., Hansen, L. P., Lützhùft, H. C., and 

Halling-Sùrensen, B. (2000): Environmental risk assessment of 

human pharmaceuticals in Denmark after normal therapeutic use. 


Suan, D. T., and Dmitrenko, L. V. (1994a): Effect of the structure of 

sulfocationites on the sorption of the antibiotic tetracycline. Appl. 

Biochem. Microbiol. 30, 629±633. 

Suan, D. T., and Dmitrenko, L. V. (1994b): Sorption kinetics of the 

antibiotic oxytetracycline on ion exchange materials. Appl. Envir. 

Microbiol. 30, 634±636. 

Süûmuth, R., Eberspächer, J., Haag, R., and Springer, W. (1987): 

Biochemisch-mikrobiologisches Praktikum. Georg Thieme Verlag, 

Stuttgart, p. 409. 

Syracuse Research Corporation (2001): Environmental Fate Data 

Base. http://esc.syrres.com/efdb.htm.. 

Thiele, S. (2000): Adsorption of the antibiotic pharmaceutical 

compound sulfapyridine by a long-term differently fertilized loess 

Chernozem. J. Plant Nutr. Soil Sci. 163, 589±594. 

Thiele, S., and Beck, I.-C. (2001): Wirkungen pharmazeutischer 

Antibiotika auf die Bodenmikroflora ± Bestimmung mittels 

ausgewählter bodenbiologischer Testverfahren. Mitteilgn. Dtsch. 

Bodenkundl. Gesellsch. 96, 383±384. 

Thiele, S., and Leinweber, P. (2000): Bedeutung der Huminstoffe für 

Bindung und Umsatz organischer Fremdstoffe ± am Beispiel 

ausgewählter Veterinärantibiotika. Rostocker Agr. Umweltwiss. 

Beitr. 8, 265±273. 

Thiele, S., Seibicke, T., and Leinweber, P. (2002): Sorption of 

sulfonamide antibiotic pharmaceuticals in soil particle size 

fractions. SETAC Europe 12th Annual Meeting, 12 ± 16 May 

2002, Vienna. 

Thiele-Bruhn, S., and Aust, M. O. (2003): Effects of slurry 

amendment on soil sorption of sulfonamide antibiotics, in 

preparation. 

Thiele-Bruhn, S., Mogk, A., and Freitag, D. (2003a): Einsatz von 

Tierarzneimitteln zur Anwendung bei landwirtschaftlichen Nutztieren 

in Mecklenburg-Vorpommern. Ber. Landwirtsch., submitted. 

Thiele-Bruhn, S., Peters, D., Halling-Sùrensen, B., and Leinweber, P. 

(2003b): Photodegradation and ageing of antibiotic pharmaceuticals 

on soil surfaces. ENVIRPHARMA, European Conference on 

Human and Veterinary Pharmaceuticlas in the Environment, 14 ± 

16 April 2003. 

Thomashow, L. S., Bonsall, R. F., and Weller, D. M. (1997): Antibiotic 

production by soil and rhizosphere microbes in situ. In C. J. Hurst, 

G. R. Knudson, M. J. McInerney, L. D. Stetzenbach and M. V. 

Walter: Manual of environmental microbiology. ASM Press, 

Washington, D.C., pp. 493±499. 

Tjùrnelund, J., Jensen, R. B., Ma, H. P., and Halling-Sùrensen, B. 

(2000): Sorption of chlortetracycline and its decomposition 

products on soil. Dansk Kemi 81, 12±14. 

Tolls, J. (2001): Sorption of veterinary pharmaceuticals in soils: a 

review. Environ. Sci. Technol. 35, 3397±3406. 

Tolls, J., Gebbink, W., and Cavallo, R. (2002): pH-dependence of 

sulfonamide antibiotic sorption: data and model evaluation. SETAC 

Europe 12th Annual Meeting 12 ± 16 May 2002, Vienna. 

Tomlinson, G., Albuquerque, C. A., and Woods, R. A. (1985): The 

effects of amidantel (Bay D-8815) and its deacylated derivative 

(Bay D-9216) on Caenorhabditis-Elegans. Europ. J. Pharmacol. 

113, 255±262. 

Topp, W. (1981): Biologie der Bodenorganismen. Quelle & Meier ± 

UTB, Heidelberg, p. 224. 

Ungemach, F. R. (2000): Figures on quantities of antibacterials used 

for different purposes in the EU countries and interpretation. Acta 

Vet. Scand. Suppl 93, 89±97.


van den Bogaard, A. E., London, N., and Stobberingh, E. E. (2000): 

Antimicrobial resistance in pig faecal samples from the Netherlands 

(five abattoirs) and Sweden. J. Antimicrob. Chemother. 45, 

663±671. 

van Dijck, P., and van de Voorde, H. (1976): Sensitivity of 

environmental microorganisms to antimicrobial agents. Appl. 

Environ. Microbiol. 31, 332±336. 

van Gool, S. (1993): Possible effects on the environment of antibiotic 

residues in animal manure. Tijdschr. Diergeneeskd. 118, 8±10. 

Warman, P. R. (1980): The effect of amprolium and aureomycin 

[antibiotic] on the nitrification of poultry manure-amended soil. Soil. 

Sci. Soc. Am. J. 44, 1333±1334. 

Watts, C. D., Crathorne, B., Fielding, M., and Killops, S. D. (1982): 

Non-volatile organic-compounds in treated waters. Environ. Health 

Persp. 46, 87±99. 

Weerasinghe, C. A., and Towner, D. (1997): Aerobic biodegradation 

of virginiamycin in soil. Environ. Toxicol. Chem. 16, 1873±1876. 

Wegener, H. C., Aarestrup, F. M., Jensen, L. B., Hammerum, A. M., 

and Bager, F. (1998): The association between the use of 

antimicrobial growth promoters and development of resistance in 

pathogenic bacteria towards growth promoting and therapeutic 

antimicrobials. J. Animal Feed Sci. 7, 7±14. 

Weimann, A., and Bojesen, G. (1999): Analysis of tetracyclines in raw 

urine by column-switching high-performance liquid chromatography 

and tandem mass spectrometry. J. Chromatogr. B 721, 

47±54. 

Wessels, J. M., Ford, W. E., Szymczak, W., and Schneider, S. (1998): 

The complexation of tetracycline and anhydrotetracycline with 

Mg2+ and Ca2+ : a spectroscopic study. J. Phys. Chem. B 102, 

9323±9331. 

Wetzstein, H. G. (2001): Chinolone in der Umwelt: Biologische 

Abbaubarkeit der Gyrasehemmer. Pharmazie in unserer Zeit 30, 

450±457. 

Wetzstein, H. G., Schmeer, N., and Karl, W. (1997): Degradation of 

the fluoroquinolone enrofloxacin by the brown rot fungus 

Gloeophyllum striatum: identification of metabolites. Appl. Environ. 

Microbiol. 63, 4272±4281. 

Wetzstein, H. G., Stadler, M., Tichy, H. V., Dalhoff, A., and Karl, W. 

(1999): Degradation of ciprofloxacin by basidiomycetes and 

identification of metabolites generated by the brown rot fungus 

Gloeophyllum striatum. Appl. Environ. Microbiol. 65, 1556±1563. 

Wetzstein, H. G., Karl, W., Hallenbach, W., Himmler, T., and 

Petersen, U. (2000): Residual antibacterial activity of metabolites 

derived from the veterinary fluroquinolone enrofloxacin. 100th 

General Meeting of the American Society for Microbiology, Los 

Angeles, CA. 

Wetzstein, H. G., Schneider, S., and Karl, W. (2002): Kinetics of the 

biotransformation of enrofloxacin in aging cattle dung. 102nd 

General Meeting of the American Society for Microbiology, Salt 

Lake City, UT. 

Williams, S. T. (1982): Are antibiotics produced in soil? Pedobiologia 

23, 427±435. 

Winckler, C., and Grafe, A. (2000): Stoffeintrag durch Tierarzneimittel 

und pharmakologisch wirksame Futterzusatzstoffe unter besonderer 

Berücksichtigung von Tetrazyklinen. UBA-Texte 44/00, 

Berlin, p. 145. 

Winckler, C., and Grafe, A. (2001): Use of veterinary drugs in 

intensive animal production: evidence for persistence of tetracycline 

in pig slurry. J. Soils Sed. 1, 66±70. 

Wollenberger, L., Halling-Sùrensen, B., and Kusk, K. O. (2000): 

Acute and chronic toxicity of veterinary antibiotics to Daphnia 

magna. Chemosphere 40, 723±730. 

Yeager, R. L., and Halley, B. A. (1990): Sorption/desorption of 

[ 14C]efrotomycin with soils. J. Agric. Food Chem. 38, 883±886. 

Zhu, J., Snow, D. D., Cassada, D. A., Monson, S. J., and Spalding, R. 

F. (2001): Analysis of oxytetracycline, tetracycline, and chlortetracycline 

in water using solid-phase extraction and liquid 

chromatography-tandem mass spectrometry. J. Chromatogr. A 

928, 177±186. 

Zuccato, E., Bagnati, R., Fioretti, F., Natangelo, M., Calamari, D., and 

Fanelli, R. (2001): Environmental loads and detection of 

pharmaceuticals in Italy, in K. Kümmerer: Pharmaceuticals in the 

Environment ± Sources, Fate, Effects and Risks. Springer, Berlin, 

pp. 19±27.

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