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Freshwater environments as reservoirs of antibiotic resistant bacteria and their role in the dissemination of antibiotic resistance genes
Journal Pre-proof Freshwater environments as reservoirs of antibiotic resistant bacteria and their role in the dissemination of antibiotic resistance genes Chika F. Nnadozie, Oghenekaro Nelson Odume PII:
S0269-7491(19)31443-5
DOI:
https://doi.org/10.1016/j.envpol.2019.113067
Reference:
ENPO 113067
To appear in:
Environmental Pollution
Received Date: 19 March 2019 Revised Date:
13 August 2019
Accepted Date: 15 August 2019
Please cite this article as: Nnadozie, C.F., Odume, O.N., Freshwater environments as reservoirs of antibiotic resistant bacteria and their role in the dissemination of antibiotic resistance genes, Environmental Pollution (2019), doi: https://doi.org/10.1016/j.envpol.2019.113067. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.
Sunlight degradation fate dilution Freshwater pool of ARB and ARGs
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Freshwater environments as reservoirs of antibiotic resistant bacteria and their role in the dissemination of antibiotic resistance genes Chika F. Nnadozie* and Oghenekaro Nelson Odume
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Unilever Centre for Environmental Water Quality, Institute for Water Research, Rhodes University, PO Box 94, Grahamstown 6140, South Africa Corresponding author*: E-mail: [email protected]; Tel.: +27 738052544
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Abstract
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Freshwater environments are susceptible to possible contamination by residual antibiotics that
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are released through different sources, such as agricultural runoffs, sewage discharges and
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leaching from nearby farms. Freshwater environment can thus become reservoirs where an
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antibiotic impact microorganisms, and is an important public health concern. Degradation and
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dilution processes fundamental for predicting the actual risk of antibiotic resistance
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dissemination from freshwater reservoirs. This study reviews major approaches for detecting
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and quantifying antibiotic resistance bacteria (ARBs) and genes (ARGs) in freshwater and their
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prevalence in these environments. Finally, the role of dilution, degradation, transmission and
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the persistence and fate of ARB/ARG in these environments are also reviewed. Culture-based
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single strain approaches and molecular techniques that include polymerase chain reaction
19
(PCR), quantitative polymerase chain reaction (qPCR) and metagenomics are techniques for
20
quantifying ARB and ARGs in freshwater environments. The level of ARBs is extremely high in
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most of the river systems (up to 98% of the total detected bacteria), followed by lakes (up to
22
77% of the total detected bacteria), compared to dam, pond, and spring (< 1%). Of most concern
23
is the occurrence of extended-spectrum β-lactamase producing Enterobacteriaceae, methicillin
24
resistant Staphylococcus aureus (MRSA) and vancomycin resistant Enterococcus (VRE), which
25
cause highly epidemic infections. Dilution and natural degradation do not completely eradicate
26
ARBs and ARGs in the freshwater environment. Even if the ARBs in freshwater are effectively
27
inactivated by sunlight, their ARG-containing DNA can still be intact and capable of transferring
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resistance to non-resistant strains. Antibiotic resistance persists and is preserved in freshwater
29
bodies polluted with high concentrations of antibiotics. Direct transmission of indigenous
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freshwater ARBs to humans as well as their transitory insertion in the microbiota can occur.
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These findings are disturbing especially for people that rely on freshwater resources for
32
drinking, crop irrigation, and food in form of fish. 1
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Introduction
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The continual overuse of antibiotics is causing their release into the environment. The release of
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antibiotics into the environment is a concern because of the consequent development of
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antibiotic resistance genes and bacteria that lead to the reduced effect of antibiotics on human
38
and animal pathogens. The prevalence of antibiotics in the environment is of growing global
39
concern both for the public and research community. Freshwater environments are among the
40
natural environments that are susceptible to possible contamination with antibiotics released
41
through different sources, such as agricultural runoffs, sewage discharges and leaching from
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nearby farms. The presence of antibiotics in the environment, a higher density of active bacteria
43
community indigenous to freshwater, creates an environment that is suitable for the
44
development of antibiotic resistance genes (Pereira et al. 2013; Marti et al. 2014; Chen et al.
45
2017). Freshwater ecosystems have thus become hotspots for horizontal gene transfer (HGT) of
46
antibiotic resistance genes, and consequently where resistance evolution occurs. The potential
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for ARBs and ARGs to persist in freshwater environments can lead to increased risk of infections
48
with resistant pathogens. ARGs can persist in the environment and ultimately return to human
49
as well as animal pathogens. The same freshwater body that is receiving wastewater also serves
50
as a source of a drinking water reservoir and even recreation (Baquero et al. 2008). In addition,
51
the spread of ARB and ARG can result in high ARG pool in environmental bacteria, therefore
52
encouraging the transfer of resistance into well-known as well as emerging pathogens (Czekalski
53
et al. 2012).
54
Investigations on the occurrences and type of antibiotic resistance in freshwater environments
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is limited compared to clinical settings, such as common indicator organisms (Enterococci and
56
coliforms), bacterial pathogens that cause infections that are difficult to treat in humans (Araújo
57
et al. 2010; Adelowo et al. 2014; Ndlovu et al. 2015; Berendonk et al. 2015; Chen et al. 2017).
58
The limited investigation of antibiotics in freshwater environments could be because antibiotic
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concentrations in these environments are presumed to be naturally low in Freshwater
60
environments. However, even if the antibiotic concentration in freshwater environments is low,
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the selection of resistant bacteria can occur (Kümmerer 2009; Marti et al. 2014). The majority
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of studies on the occurrence and fate of ARBs and ARGs in freshwater environments, including 2
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streams, lakes and rivers, are focussed mostly on discharge point of wastewater treatment
64
plants (WWTPs) (LaPara et al. 2011; Marti et al. 2013; Chen et al. 2013; Sabri et al. 2018).
65
Concentrations of ARBs and ARGs are usually high at the point of effluent discharges from
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WWTWP into streams, but gradually become reduced downstream of the discharge point
67
(LaPara et al. 2011; Marti et al. 2013; Chen et al. 2013; Sabri et al. 2018). The decline in
68
concentration of ARGs and ARBs downstream of WWTW effluent discharge points can be as a
69
result of different factors such as dilution, degradation, adsorption, transport (Floehr et al. 2013;
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Anyaduba 2016; Nelson et al. 2018; Yoon et al. 2018). Among these different factors, dilution
71
and degradation play a significant role in the fate of ARGs in freshwater environments (Snow et
72
al. 2015; LaPara et al. 2015; Anyaduba 2016; Jerde et al. 2016; Collins et al. 2018) (Figure 1).
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Thus, this paper investigates how dilution and degradation influence the fate of ARB, and ARG,
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and the subsequent development and dissemination of resistance to human pathogens. Before
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proceeding to evaluate the influences of dilution and degradation on ARB and ARG, we first
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present an analytic review of techniques for detecting and quantifying ARBs and ARGs in
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freshwater environments. The abundance and diversity of resistance genes in freshwater
78
environments are also reviewed.
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Techniques for quantifying antibiotic resistance genes and bacteria in freshwater bodies
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Culture-based single strain approaches are the most frequently applied techniques for
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quantifying ARB and ARGs in freshwater environments (West et al. 2010; Suzuki et al. 2013;
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Zurfluh et al. 2013). Culture-based approaches involve growing the microorganisms on a
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nutrient medium (most preferably a growth medium that is selective to the bacteria of interest).
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Individual isolates are tested for antibiotic resistance by growing them on nutrient medium
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supplemented with the antibiotics of interest (West et al. 2010), or embedded within an
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antibiotic disk (Narciso-da-Rocha and Manaia 2016). In most cases, the antibiotics are those
88
commonly used in human as well as veterinary medicine that show different mechanisms and
89
pathways for activity, such as inhibition of cell wall synthesis or protein synthesis (Moore 2013).
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The culture-based approach is limited because it does not capture the major components of the
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natural assemblage of a freshwater microbial community, and ARGs detected through such a
92
method is just a selection of the total bacteria community (Suzuki et al. 2013). Alternative
3
93
approaches involve the application of molecular tools , including Polymerase chain reaction and
94
quantitative polymerase chain reaction (PCR and qPCR) (Marti and Balcázar 2013; Suzuki et al.
95
2013; Mao et al. 2014; Czekalski et al. 2015; Xiong et al. 2015; Di Cesare et al. 2017; Zhou et al.
96
2017; Giebułtowicz et al. 2018; Sabri et al. 2018; Zheng et al. 2018; Liu et al. 2019), and
97
metagenomics profiling ARB and ARG of the environment (Kristiansson et al. 2011; Port et al.
98
2012; Amos et al. 2014, 2018; Bengtsson-Palme et al. 2014; Xiong et al. 2015; Fitzpatrick and
99
Walsh 2016; Zheng et al. 2017; Fresia et al. 2018).
100
Polymerase chain reaction (PCR) uses small quantity of a sample DNA (template), two primers
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flanking the target sequence (ARG and ARB), nucleotides, together with thermostable DNA
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polymerase to amplify a particular region of DNA of interest, creating a large quantity of DNA
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from a very small environmental sample (Boyle 2014). qPCR is an improvement of the
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conventional PCR ,of which the targeted gene is amplified and quantified at the same time
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(Pabinger et al. 2014). Quantification of the number of the target present in template DNA is
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extremely challenging with conventional PCR. With qPCR, the quantity of the product that is
107
produced is observed in the course of the reaction by observing the fluorescence of dyes or
108
probes that are added in the reaction mix. The amount of fluorescence of dyes or probes is
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proportional to the amount of product formed, and the amount of amplification cycles
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necessary for obtaining a particular number of DNA molecules is recorded. By assuming a
111
certain amplification efficiency, one can calculate the amount of DNA molecule of the amplified
112
target sequence that is originally present in the sample (Kubista et al. 2006). The lesser the PCR
113
cycles that is needed to make enough material for detection, the more the copies of DNA
114
template that is present at the beginning of the experiments (Pabinger et al. 2014)
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Metagenomics involves characterizing the total microbial community DNA (which includes
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prokaryotes, viruses, and eukaryotes) in an environment (Handelsman et al. 1998; Staley and
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Sadowsky 2016). Thus, using a single dataset of DNA sequence, one can quantify all known
118
resistance genes present in the sample (Munk et al. 2017). Metagenomics can be through
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sequence-based or function based screening. A sequence-based metagenomic uses sequencing-
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by-synthesis technology to produce millions of sequence reads without the need for cloning
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(Gilbert and Dupont 2011). Screening through this approach does not require prior knowledge
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of the sequence of interest. Alternatively, a function-based screening is a method whereby DNA 4
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clones of the environmental library are screened for their ability to confer a function of interest
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to a heterologous host. The advantage of this approach is that it does not require prior
125
knowledge of the sequence. Therefore, it provides direct evidence of the function of genes that
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are not previously known (Martínez and Osburne 2013). Functional screening, whereby fosmid
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libraries were constructed and screened to evaluate the level of antibiotic resistance have been
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undertaken for river water and sediments (Amos et al. 2014; González-Plaza et al. 2018).
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Compared to a sequence-based metagenomic screening, a function-based screening requires
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less amount of DNA from environmental samples (Staley et al. 2015).
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Metagenomics offers advantages over PCR because it is possible to quantify thousands of target
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genes in environmental samples without prior knowledge of bacteria or genes present in the
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sample. However, this approach is limited in several ways: i) determining the function of newly
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discovered sequences can be a challenge, ii) large fractions of the metagenomic sequences
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generated usually do not show any significant similarity to previously annotated sequences, thus
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making annotation difficult (Nnadozie et al. 2017).
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Prevalence of antibiotic resistant bacteria in freshwater ecosystems
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Freshwater environments are an important reservoir of ARBs and ARGs. Table 1 reflects the
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prevalence of ARBs in freshwater ecosystems. One can observe that they are widespread in
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different types of freshwater environments. The level of ARBs is extremely high in most of the
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river systems (up to 98% of the total detected bacteria), followed by lakes (up to 77% of the
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total detected bacteria). This indicates that rivers and lakes serve as a significant reservoir for
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the spread of antibiotic resistance to opportunistic pathogens. The high concentration of ARBs
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that are resistant to cephalothin, penicillin, tetracycline, ampicillin, and chloramphenicol in
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rivers and lakes, indicate that these antibiotics are present in the freshwater systems. This could
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be because of the widespread use of these antibiotics as basic antimicrobial drugs. The
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widespread application of antibiotics leads to a high proportion of bacteria are resistant to
149
them. A number of factors including the geographical location of the region and prescription
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policies can influence the proportion of ARBs in freshwater environments. In regions where
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restrictive rules exist regarding prescription and disposal, a low prevalence of ARB has been
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reported (Tacão et al. 2015). 5
153
On the other hand, the concentration of ARBs in the dam, pond, and spring (< 1%) are relatively
154
low (Table1). Yet, dam and pond systems could be underappreciated reservoir s of ARBs. The
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prevalence of ARBs could be high in these environments, considering that usually only a sample
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is withdrawn during sampling. The reason for the higher frequency of ARBs in river and lake
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compared to the dam and pond require further investigation.
158
Of most concern to public health is the occurrence of extended-spectrum β-lactamase
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producing Enterobacteriaceae, Carbapenemase-producing Enterobacteriaceae, methicillin
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resistant Staphylococcus aureus (MRSA) and vancomycin resistant Enterococcus (VRE) – all of
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which can cause potentially highly epidemic infections (Arora et al. 2014) (Table 2). A worldwide
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environmental dissemination of these ARBs hosting ARGs has been suggested in recent studies
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(Spindler et al. 2012; Zhang et al. 2013; Zurfluh et al. 2013; Khan et al. 2013; Czekalski et al.
164
2014; Laht et al. 2014; Devarajan et al. 2015, 2016; Djenadi 2017). Their presence in freshwater
165
environments has been reported in countries around the world, such as Portugal, Finland,
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France, Pakistan, Switzerland, United States of America, China and Brazil (Spindler et al. 2012;
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Zhang et al. 2013; Zurfluh et al. 2013; Khan et al. 2013; Czekalski et al. 2014; Laht et al. 2014;
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Devarajan et al. 2015, 2016; Djenadi 2017).
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The family Enterobacteriaceae are Gram-negative bacteria, including Escherichia coli (E. coli),
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Citrobacter, Salmonella, Yersinia pestis, Shigella, Proteus, Enterobacter, Salmonella and
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Klebsiella (Patel and Nagel 2015). Extended-spectrum β-lactamase and Carbapenemase-
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producing Enterobacteriaceae (CPE), such as E. coli, K. pneumoniae, E. cloacae, C. frendii,
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Enterobacter asburiae and Klebsiella oxytoca isolates have been detected in freshwater
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environments such as streams and rivers (Table 2) (Zurfluh et al. 2013; Ye et al. 2017; Harmon et
175
al. 2019). The β-lactamase-producing Enterobacteriaceae are resistant to the modern extended-
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spectrum cephalosporins antibiotics as well as penicillins and monobactams (Bush and Jacoby
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2010; Miyagi and Hirai 2019). Their resistance to these antibiotics is due to their plasmid-
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mediated production of enzymes that hydrolyze the β-lactam ring of the antibiotic compounds.
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This type of resistance is caused by a high number of point mutation variations of established
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broad spectrum β –lactamases – the so-called extended-spectrum β-lactamases (ESBLs). Several
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ESBLs are members of SHV and TEM β-lactamases, while others are classified as OXA, CTX-M,
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GES, TLA and PER β-lactamases (Poirel et al. 2000; Coque et al. 2008; Bush and Jacoby 2010). 6
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The reported high occurrence of ESBL in Zurfluh et al. (2013) is disturbing because the study was
184
carried out in Switzerland where strict policies apply to antibiotic use (Filippini et al. 2006;
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Zurfluh et al. 2013). The identification of ESBL genes in environmental Aeromonas,
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Pseudomonas sp., and Acinetobacter isolates revealed an extensive assortment of mobilization
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events, implying that these bacteria are vehicles for ESBL dissemination (Girlich et al. 2011;
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Tacão et al. 2012). Although the number of studies exploring freshwater as possible reservoirs of
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ESBL producers is still scarce, their reported presence in the freshwater environments is highly
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disturbing because of the potential to transmit their genes into human pathogens (Flores
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Ribeiro et al. 2012; Zhou et al. 2014; Czekalski et al. 2015; Chen et al. 2015, 2017).
192
Furthermore, carbapenemases are a large group of β-lactamases that are grouped into classes
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ABC and D that inactivate carbapenem antibiotics. Carbapenem antibiotics, such as imipenem
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and meropenem are used as last resort drugs to treat dangerous infections caused by ESBL
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producers. Due to the limited antibiotic options, the treatment of infections caused by the
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organisms that are resistant to carbapenem present serious challenges (Izadpanah and Khalili
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2015). CPE has been designated as urgent threats (Centers for Disease Control and Prevention
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2013, 2018) and is associated with very high mortality rates. Therefore the presence of CPE in
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the freshwater bodies presents a huge challenge to public health (Harmon et al. 2019). Table 2
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shows that the CPE including E. coli, K. pneumoniae, K. oxytoca and E. cloacae that are positive
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for KPC-2, VIM- and IMI-2 were recovered in the bottom sediments of freshwater environments.
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Increased loading and persistence have been implicated for their recoveries in the sediments
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(Piedra-Carrasco et al. 2017). The IMI and KPC enzymes belong to Class A, which are inhibited by
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clavulanic acid. The KPC enzymes are plasmid coded. The VIM types belong to Class B
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carbapenemases and are integron coded (Bush and Jacoby 2010; Zurfluh et al. 2013). The
206
increased loading and persistence of CPE implies that these ARBs are detectable in sediments,
207
even if they are not found in water columns. Klebsiella and Enterobacter seemed to be the most
208
prevalent microorganism that produces KPC-2 (Picão et al. 2013)
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The occurrence of methicillin-resistant Staphylococcus aureus (MRSA) in the recreational
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freshwater environment suggests a potential colonization of people that come in contact with
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them, and possible environmental contamination (Levin-Edens et al. 2012; Fogarty et al. 2015;
212
Hatcher et al. 2016; Thapaliya et al. 2017). Infections caused by MRSA are difficult to treat. 7
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Staphylococcus strains that carry methicillin resistance gene mecA on their staphylococcal
214
cassette chromosome of type II cause healthcare-associated MRSA infections. They are resistant
215
to many classes of antibiotics such as oxacillin methicillin, penicillin, and amoxicillin.
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Alternatively, Staphylococcus strains that carry Panton-Valentine leukocidin (PVL) genes on their
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SCCmec type IV cause community-associated MRSA (CA-MRSA) infections, and are resistant to
218
fewer classes of antibiotics.
219
In addition, antimicrobial–resistant Enterococci spp. isolates have been recovered in freshwater
220
environments (Table 2). For example, vanCtype VRE is widely distributed in aquatic
221
environments, including rivers and coastal areas (Zdragas et al. 2008; Nam et al. 2013;
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Nishiyama et al. 2017). Vancomycin used to be the most effective last-line-of-defense
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antimicrobials for managing antimicrobial resistant Enterococci, but its efficiency is now in doubt
224
(Arias and Murray 2012). vanA, vanB, vanC, vanD, vanE and vanG are the six recognised
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vancomycin-resistant genes (Khan et al. 2008). vanA genes are associated with non-conjugative
226
and conjugative plasmids carrying Tn1546‐like transposons, whereas vanB genes are associated
227
with conjugative transposons, which includes Tn5382, Tn1549 and Tn1547 (Roberts et al. 2009).
228
At this juncture, it is noteworthy that the levels of ARGs presented in this section are based on
229
either total heterotrophic bacteria counts or particular pathogenic bacteria cultured in the
230
laboratory and must be viewed as representing a fraction of bacteria that actually occur in
231
freshwater environments.
232 233
The abundance of antibiotic resistance genes in freshwater environments detected using non-culture based techniques
234
Using
235
chloramphenicol as well as Macrolide-Lincosamide-Streptogramin resistance genes have been
236
widely reported in freshwater bodies (Amy Pruden et al. 2006; Pei et al. 2006; Storteboom et al.
237
2010; Czekalski et al. 2012, 2014, 2015; Szekeres et al. 2018) (Table 3). The majority of studies
238
have focused on sulfonamide resistance genes (sul1 and sul2) compared to the other resistance
239
genes. The extensive focus on sulfonamide resistance genes could be attributed to the fact that
240
they are considered as important indicators of freshwater pollution. Sulfonamide resistance
241
genes are less abundant in pristine environments. sul1 abundance is usually associated with
242
input from WWTPs effluent discharges into freshwater, while sul2 is mainly associated with
molecular
techniques,
sulfonamides,
8
tetracyclines,
β-lactam,
aminoglycosides,
243
inputs from urban activities such as sand dredging, agricultural runoff, urban discharges,
244
religious rituals, open defecation, and other anthropogenic activities (Devarajan et al. 2016). The
245
presence of sul genes in the freshwater environment is not driven by the commonly known
246
drivers such as nutrient enrichment, pH, and the quality and amount of organic carbon of the
247
compositional bacterial community in the freshwater environment. Czekalski et al. (2015)
248
provide a baseline useful for distinguishing between pollution-induced concentrations of ARG
249
from the natural background levels. Freshwater bodies with sul1 and sul2 abundances of
250
2.8 × 10− 3 ± 3.9 × 10− 4 and 3.1 × 10− 3 ± 3.1 × 10− 3 are considered not impacted by ARG pollution
251
(Czekalski et al. 2015). Thus, sul1 and sul2 abundances ranging 1.5 × 10− 3 to 1.6 × 10− 2 and
252
3.1 × 10− 3–7.2 × 10− 5, respectively are considered baseline. The values provided are ratios of
253
both sul genes normalized to eubacterial 16S rRNA genes.
254
ARGs are considered contaminants of emerging concern because some of the genes are present
255
in bacteria as structurally innate reservoir genes. blaTEM genes are usually obtainable in most
256
environments regardless of whether the site has been exposed to anthropogenic contamination
257
and this has been attributed to their ubiquitous presence as housekeeping genes (Demaneche
258
et al. 2008). aadA and blaTEM genes have also been reported to be present even before the start
259
of the twentieth century (Devarajan et al. 2015). blaCTX-M gene is the most common ESBL
260
globally, with CTX-M-15 and CTX-M-14 mainly invading humans, animals and the environment
261
(Cantón et al. 2012).
262
It is important to note that resistance genes in Table 3 do not necessarily represent the full
263
range of genes freshwater bodies as other ARGs may occur. It is possible that the abundance of
264
other resistance genes is not sufficient enough to be captured by the method applied. The
265
studies in Table 3 applied qPCR methods, which are limited. Reported variations in the
266
abundance of resistance genes could be due to a number of factors including the period of
267
sampling, disposable practices, and sampling season (Devarajan et al. 2016). Studies on factors
268
influencing the variation of abundance of resistance genes in the environment are scarce but
269
necessary for establishing a reliable threshold for analyzing ARG pollution.
9
272
Effects of dilution and degradation on the fate of antibiotic resistant bacteria and antibiotic resistance genes in the freshwater environment
273
The previous section discusses factors that drive the persistence of antibiotic resistance in
274
Freshwater without regarding the issues of degradation, dilution, transportation, and
275
adsorption. These processes are particularly important in making ecologically relevant
276
inferences due to the impacts of water volumes, the presence of substrate surfaces, current and
277
tides (Foote et al. 2012; Jerde et al. 2016; Shogren et al. 2017). Antibiotic resistant genes and
278
antibiotic resistant bacteria in aquatic environments can exist either as intracellularly in viable
279
antibiotic resistant bacteria as genomic and plasmid DNA or extracellularly as free eDNA
280
(environmental DNA) that is shielded within phage capsids, extracellular polymeric substances
281
(EPS), cell debris or on clay mineral surfaces (Dodd 2012). In terms of their fate, the ARBs and
282
ARGs can be transported, experience decay through biological and non-biological processes,
283
adsorbed onto particulate matter, diluted, uptake by aquatic microorganisms through HGT
284
(Chen and Dubnau 2004; LaPara et al. 2015; Anyaduba 2016). Transport means the movement
285
of water by advection or diffusion in streams or ponds, and ARGs and ARBs can be transported
286
over considerable distances in water (Poté et al. 2003). Compared to the other processes,
287
transport does not play a substantial role in the fate of ARBs and ARGS in aquatic environments
288
(Anyaduba 2016). Studies suggest that it is dependent on the flow regime of the water and
289
generally in high flowing rivers the ARBs and ARGs are diluted, and so it is more of dilution effect
290
(LaPara et al. 2015; Jerde et al. 2016). This explains why rivers and lakes located in areas prone
291
to drought are expected to have a greater presence of ARBs and ARGs (Anyaduba 2016).
292
Transport will mainly influence the representativeness of the concentration of ARBs and ARG
293
that is measured at one point of a Freshwater environment (Goldberg et al. 2016). On the other
294
hand, adsorption is the deposition and trapping in substrate crevices, sticking to stream biofilms
295
colonizing substrate surfaces (Anyaduba 2016; Jerde et al. 2016; Shogren et al. 2017). ARGs
296
interact with minerals, humid acid in the freshwater environment, explaining why the majority
297
of ARGs remain immobilized in sediments. Immobilization of ARGs will lead to longer
298
persistence in water (Ficetola et al. 2008). The extended persistence of ARGs by adsorption
299
creates an opportunity for resuspension to occur, for instance when there is shear stress or high
300
flow events (Turner et al. 2015).
270 271
10
301
Degradation and dilution have been found to play major significant role in the control of the fate
302
and persistence of microbial contaminants in Freshwater and are fundamental for predicting the
303
actual risk of antibiotic resistance dissemination from freshwater reservoirs (Foote et al. 2012;
304
Lasagna et al. 2013; Anyaduba 2016; Goldberg et al. 2016). Therefore the following section
305
discusses how degradation and dilution determine the persistence of ARBs and ARGs in
306
freshwater.
307
Degradation
308
Degradation is the most significant process that influences the fate of ARBs and ARGs in the
309
freshwater environment (Goldberg et al. 2016). Degradation can occur due to either microbial
310
activity or natural sunlight exposure.
311
microbiome secretes enzymes during metabolism to break down large organic molecules before
312
they are assimilated into cells. Thus, environmental ARGs are prone to the biotic degradation
313
process that involves DNAses. DNAses catalyze the cleaving of the phosphodiester bond
314
between the phosphate group and the deoxyribose sugars in the DNA, releasing nucleotides
315
that are assimilated by the organisms. Studies have demonstrated DNA degradation in aquatic
316
systems by high microbial and enzymatic activity (Zhu 2006; Dejean et al. 2011; Pilliod et al.
317
2014). ARGs discharged from wastewater treatment facilities into receiving water significantly
318
have a loss of mechanisms due to natural decay (Anyaduba 2016; Jerde et al. 2016; Shogren et
319
al. 2017). Higher degradation rates occur with environmental conditions, such as neutral pH,
320
moderately high temperature and UV-B irradiation concomitantly that favor microbial growth
321
and increase the presence of nucleases, which enhance ARG loss in ecosystems (Strickler et al.
322
2015; Anyaduba 2016). However, sediments provide ARGs protection from biological
323
degradation (Turner et al. 2015). In previous studies, DNA protection from degradation upon
324
adsorption onto the sediments has been explained in different ways. DNA is protected from
325
nucleases on adsorption onto sediments (Torti et al. 2015). It is suggested that adsorption onto
326
sediments reduces accessibility to nucleases (Khanna and Stotzky 1992). Nucleases adsorbed on
327
sediment become inactivated (Lorenz and Wackernagel 1987, 1992; Sarkar et al. 1989;
328
Romanowski et al. 1991; Paget et al. 1992). The DNA can also bind to biogenic sediment
329
components, which include proteins and humic acids that can enhance the resistance to
With regards to microbial activity, organotrophic
11
330
nucleases (Nielsen et al. 2006). Furthermore, the DNA may interact with exopolymeric
331
substances of sediment particles that will shield them from degradation.
332
In the case of natural sunlight, it is a generally accepted disinfectant and has always been relied
333
on as a strong determinant of the persistence of microbial pollutants in surface waters, including
334
fresh and marine surface waters (Boyle et al. 2008; Nelson et al. 2018). Sunlight degradation of
335
bacteria is by photoinactivation and is well studied. Sunlight degradation has been observed in
336
Salmonella enterica, Shigella flexneri, Escherichia coli (E. coli) and Enterococcus faecalis (E.
337
faecalis) after sunlight exposure (Berney et al. 2007; Sassoubre et al. 2014; Mcclary et al. 2017;
338
Scoullos et al. 2019; Busse et al. 2019). More specifically, few studies exist that demonstrate
339
sunlight mediated degradation of bacteria in freshwater.
340
coliforms, E. coli, Legionella spp. Enterococci, F-RNA, and somatic coliphages, phages was
341
observed in freshwater (Dutka 1984; Davies and Evison 1991; Sinton et al. 2002; Dick et al. 2010;
342
Korajkic et al. 2014; Wanjugi et al. 2016).
343
However, sunlight can vary substantially in spectral quality, and underwater. Apart from sunset
344
and sunrise, water particles can change the spectral composition of sunlight. Furthermore, solar
345
radiation differs seasonally (Salter 2018), with longer sunlight duration occurring during summer
346
at high latitudes, and no sunlight at all during winter. Differences in seasonal distribution,
347
strength and biological activity of UV radiation have been recognized in the literature for while
348
(Jablonski and Chaplin 2010). The geometry of sunlight that reaches different places varies in
349
different seasons. An implication of this seasonal differences in solar irradiation underwater is
350
that sunlight mediated degradation of microorganisms in water will vary based on season,
351
weather condition, time of the day and location (Nelson et al. 2018). In the study by Calero-
352
Cáceres et al. (2017), higher prevalence and abundance of ARGs were observed in winter, and
353
this was attributed to lower irradiance and temperature in winter. Besides, a greater total
354
organic carbon in water during winter has been suggested to could cause bacterial regrowth,
355
which leads to a higher abundance of ARGs compared to summer seasons (Calero-Cáceres and
356
Muniesa 2016).
357
Also, different intensities of solar UV radiation are obtainable in different regions. Therefore, in
358
some regions, the intensity of sunlight radiation may not be sufficient to initiate sunlight
12
Sunlight inactivation of faecal
359
mediated degradation. Compared to Eurasia regions, it is suggested that microorganisms in
360
waters located in regions that are near the equator, and in high altitudes experience high UVR
361
(Jablonski and Chaplin 2010; Nguyen et al. 2014). Therefore, more sunlight mediated
362
degradation is expected.
363
Additionally, factors such as cloudiness, aerosols, stratospheric ozone (Zepp et al. 2018), water
364
quality, snow and ice over, and depth at which the microorganisms reflexively circulate can
365
modify the exposure of the microorganisms to solar radiation (Sulzberger et al. 2019;
366
Williamson et al. 2019). Snow-cover on ice can inhibit most or all UV radiation from entering the
367
water column, depending on how thick it is. Regarding mixed layer depth, the exposure of
368
microorganisms to sunlight mediated degradation depends on the vertical position within the
369
water column. Microorganisms that are trapped near the surface are more exposed to sunlight
370
mediated degradation (Williamson et al. 2019).
371
Furthermore, the water quality and depth exert appreciable influence over the rate of
372
photoinactivation. The presence of dissolved organic matter in water reduces deeper
373
penetration of light (attenuation). The amount of dissolved organic matter (DOM) affects
374
exposure required sunlight mediated degradation of microorganisms. Suspended sediments,
375
presence of phytoplanktons, eutrophication exacerbates irradiance attenuation. A key
376
implication to these is that in humic-stained or eutrophic waters, sunlight mediated degradation
377
is less likely (Nelson et al. 2018).
378
Besides, bacteria have the ability to repair sunlight damage, and there is a chance for recovery
379
and regrowth if the injury by sunlight is sublethal (Nelson et al. 2018). Worse still, some
380
nosocomial pathogens (Serratia marcescens, Pseudomonas putida, and Stenotrophomonas) are
381
resistant to solar radiation (Glady-Croue et al. 2018). ARBs are more resistant to inactivation
382
during solar irradiation than non-resistant bacteria (Al-Jassim et al. 2017). Furthermore, a study
383
investigating the effect of sunlight on the decay of faecal indicator bacteria (FIB) in freshwater
384
demonstrated the minimal effect of sunlight on the survival of FIBs in freshwater (Korajkic et al.
385
2019). Bacterial photostress is complex and is determined by multiple environmental stressors.
386
For instance, higher concentrations of dissolved oxygen in the environment is required to
387
improve photoinduced damage of bacteria (Mcclary et al. 2017).
13
388
Sunlight-mediated pathogen inactivation does not guarantee a reduction in overall risk to
389
pathogen dissemination and the geographic kinds of human pathogens vary with distribution
390
and biological activity of UV radiation may change (Boehm et al. 2018).
391
Even if ARBs are effectively inactivated by sunlight, their DNA which contain ARG can still be
392
intact and capable of transferring resistance to non-resistant strains (Anyaduba 2016).
393
Information on the ability of sunlight to degrade ARGS is still scarce, but studies provide
394
evidence of monochromatic UVC radiation (UV254) preventing several ARGs from converting
395
competent non-resistant recipient bacteria into conforming resistant phenotypes (Nobuo and
396
Ikeda 1969; Setlow 1977; Chang et al. 2017). It is well established that UV can induce DNA
397
damage and leads to DNA base lesions such as pyrimidine (6-4) pyrimidone adducts [(6-4) and
398
cyclobutane-pyrimidine dimers (CPDs) photoproducts] (McKinney and Pruden 2012; Destiani et
399
al. 2018). A significant rate of UV–induced ARG degradation is documented (Destiani et al.
400
2018; Yoon et al. 2018). Also, exposure to artificial UV light eliminated plasmid-borne resistance
401
gene (Schuch and Menck 2010; Yoon et al. 2018). However, the plasmid-borne ARG that was
402
eliminated by UV is repairable during transformation by competent bacterial cells within the
403
environment (Nelson et al. 2018). The significant repair was observed in E. coli recipient strain
404
(DH5α) (Yoon et al. 2018). Also, ARGs degradation is slow (Nelson et al. 2018). Therefore,
405
sunlight mediated degradation does not completely eliminate ARGs in the freshwater
406
environment. The ARGs can be taken up by a competent bacteria, and incorporated into the
407
bacteria genome even when the original donor ARB cell is absent (Yoon et al. 2018). The rate at
408
which HGT occurs in the aquatic environment is not known quantitatively and is important to
409
determine whether HGT is a sink or source of ARGs in the aquatic environment.
410
Dilution
411
Dilution means the reduction of the concentration of a substance as it is dissolved in a larger
412
volume of a solvent (in this case water) (Lasagna et al. 2016). In aquatic environments dilution of
413
pollutants such as eDNA and microorganisms depends on currents, dynamic hydrological
414
processes and the amount of pollutants received (Floehr et al. 2013; Baldigo et al. 2017; Collins
415
et al. 2018). Where the body of water is large, with water current and strong tide the ARBs and
416
ARGs in water will quickly dilute and scatter (Foote et al. 2012). In addition, any substance
417
present in flowing water moves with the current. The dynamic hydrological processes include 14
418
advection and diffusion (Van Genuchten et al. 2013; Molz 2015). Advection refers to the
419
movement of objects suspended or dissolved in water along with the bulk flow, such as when a
420
river is flowing down a stream. Diffusion is the net motion of objects from a place of high
421
concentration to a place of low concentration, and therefore the substance spreads out in the
422
river. In terms of the amount of pollutants received, the assimilating capacity of the freshwater
423
can only cope with a certain part of the pollutants.
424
Dilution plays a substantial role in the reduction of the level of pollutants in water (Foote et al.
425
2012; Lasagna et al. 2013). It has been suggested as a process that is ever-present and not
426
influenced by either biological and chemical conditions (Lasagna et al. 2013). However, while
427
dilution reduces the concentration of a contaminant, it does not eliminate it from the system. In
428
agreement with the law of conservation of mass, the pollutants in freshwater become a part of
429
the hydrological cycle that is separated through adsorption and consequently become
430
dangerous wastes that will become of grave concern overtime(Sharma and Ahmad 2014).
431
Persistence of antibiotic resistance in the bacterial population
432
The presence of high concentrations of antibiotics in the environment causes direct selection for
433
resistance markers (Andersson and Hughes 2011). Therefore, freshwater bodies receiving
434
pharmaceutical waste streams or runoffs from agricultural farmlands can become polluted with
435
high concentrations of antibiotics, enabling selective pressure that encourages the preservation
436
and permanency of antibiotic resistance. Antibiotics, even at low concentrations can select for
437
resistant bacteria in the environment (Gullberg et al. 2011; Alm et al. 2014). Resistance becomes
438
permanent in bacterial populations under certain environmental conditions, even where
439
antibiotic selective pressure is low or absent for a significant period. Resistance can persist by (i)
440
compensatory mutations that reinstate fitness without loss of resistance; (ii) the incidence of
441
infrequent cost-free resistance mutation (iii) genetic association and co-selection between
442
resistance mutations and extra selected genetic markers (e.g. resistances or virulence factors)
443
(Andersson and Hughes 2011). Compensatory evolution lessens costs and allows maintenance
444
of resistance even without selective pressure.
445
Nevertheless, antibiotic resistance is not always the primary purpose of antibiotic resistance
446
genes. It has been suggested that some genes code for functions that confer a selective benefit 15
447
in the natural environment (Alm et al. 2014). This explained why Shwanella oneidensis MR-I
448
isolated from pristine sediments devoid of pharmaceutical impact possessed a Mex system
449
capable of antibiotic efflux, improving the fitness of the bacteria in the sediments (Groh et al.
450
2007). If the freshwater environment contains other substances, including anti-microbial
451
peptides released by other microbes, metals and organics, then Shwanella oneidensis can
452
benefit for common efflux pumps that are also applied for resistance to antibiotics (Alm et al.
453
2014). In terms of compensatory mutation, the fitness cost that is accompanying antibiotic
454
resistance is reduced over time as the bacteria become accustomed to the environment, and so
455
the pressure to lose the resistance phenotype is reduced (Moore et al. 2000). In terms of co-
456
selection, it is most likely that the resistance plasmids bear additional genes coding for proteins
457
that improve colonization, enabling alternative carbon utilization or improved nutrient uptake.
458
ARGs that have become fixed in a bacterium are difficult to eliminate (Andersson and Hughes
459
2011).
460
Therefore, having elucidated the processes that are involved in propagation and persistence of
461
antibiotic resistance in bacterial populations, as well as several mechanisms that contribute to
462
the stability of antibiotic resistance within the bacterial populations, it can be deduced that that
463
the release of antibiotics into freshwater environments contribute to the selection for resistance
464
within the bacterial population. Also important, one can deduce that antibiotic resistance
465
mechanisms will persist in freshwater systems in the presence of other substances that are not
466
antibiotics because the bacteria use mechanism that is similar for antibiotic resistance to survive
467
in the presence of substances that are toxic to them.
468 469
Investigating the potential transfer of antibiotic resistance in freshwater bodies
470
The prevalence of ARB and ARGs in freshwater environments implies that they are hotspots for
471
possible ARG dissemination. ARGs can be transferred through HGT mechanisms, including
472
transduction, conjugation, and transformation. In hotspots environments, different genetic
473
structures bearing ARGs are usually detected in addition to the ARGs themselves (Lupo et al.
474
2012). There are various types of mobile genetic elements, such as integrons, transposons,
475
plasmids, bacteriophages, as well as a combination of them. In particular, integrons are useful
476
targets for the detection of possible ARG transfer and spread because (i) they are one of the 16
477
simplest elements that participate in the transfer (mobility) of gene cassettes (ii) all integrons
478
have a common structure (iii) integrons can be linked to other mobile genetic elements, and (iv)
479
they can efficiently trap ARGs. Integrons can acquire, express and exchange ARGs that are
480
embedded in gene cassettes (GCs). An integron can be defined based on its structural
481
components. An integron is composed of 3 elements, namely intl gene, which encodes an
482
integrase, site attl specific for recombination and a promoter attI (Heuer et al. 2004; Cambray,
483
et al. 2010; Rizzo et al. 2013). Generally, GCs have an open reading frame that is coupled to an
484
attC site that is integrated or cut from the functioning platform by a site-specific recombination
485
mechanism that catalyze the intl integrase (Rizzo et al. 2013).
486
The two major groups of integrons are resistance integrons (RIs) and chromosomal integrons
487
(Cis). Cis is found on the chromosome of many bacterial species (Cambray, et al. 2010). RIs have
488
been investigated in several gram-negative bacteria, and few gram-positive bacteria (Nandi et
489
al. 2004; Xu et al. 2010). RIs are carried on mobile genetic elements, including plasmids and
490
transposons, which facilitate their transfer among bacteria. RIs are classified into five groups
491
based on the sequence of amino acid of the intl protein (Cambray, et al. 2010). Classes 1, 2 and
492
3 are the most studied. Classes 1 and 2 are frequently observed in environmental samples,
493
human and bacterial isolates (Stokes and Gillings 2011). Class 1 RIs are extensively described
494
and mostly encountered in multidrug-resistant bacteria. Only four class 3 RIs have been studied
495
in environmental Delftia isolates and clinical Enterobacteriaceaen strains (Xu et al. 2007).
496
Over 130 GCs encode resistance to nearly all antibiotics families such as aminoglycosides,
497
chloramphenicol, fosfomycin, lincosamides, β-lactams, rifampicin, macrolides, and quinolones. It
498
is because GC bore on integrons encodes resistance for a wide range of antibiotics that targeting
499
integron structures will provide an overall perspective on antibiotic resistance and the spread of
500
ARGs in the environment (Kristiansson et al. 2011). Fortunately, a qPCR method that targets
501
three main classes of integrons in DNA from complex matrices has been described (Barraud and
502
Ploy 2011). Thus, RIs can be applied in tracking the occurrence of resistance genes within the
503
environment. Furthermore, the presence of RIs can be used to indicate the acquisition of
504
antibiotic resistance genes. The prevalence of the Class 1 integron-integrase gene (intl1) has
505
been recommended as a reliable proxy for the occurrence of pollution by anthropogenic sources
506
(Gillings et al. 2015). Studies that show the occurrence of integron-bearing drug-resistant 17
507
bacteria in freshwater environments exists. In a study by Koczura et al. (2015), integron‐carrying
508
multidrug resistant coliform bacteria bearing virulence genes was observed in recreational lakes.
509
The presence of E. coli strains bearing Class 1 and 2 integrons isolated from freshwater in
510
Australia has been reported (Sidhu et al. 2017). A high prevalence of E.coli bearing class
511
1integrons was observed in the Minjian River in China (Chen et al. 2011).
512
Besides the class 1 integrase gene (intI1, transposase gene (tnpA) is another marker for mobile
513
genetic elements, that are important in the dissemination of resistance (Zhu et al. 2013; Gillings
514
et al. 2015; Hu et al. 2016). tnpA genes were detected in groundwater samples (Szekeres et al.
515
2018). However, intI1 responds quicker to environmental stressors than tnpA and so are
516
proposed as a better proxy for anthropogenic pollution and resistant (Gillings et al. 2015;
517
Szekeres et al. 2018)
518
In practice, however, it is very difficult to deduce the contribution of each and every
519
phenomenon to ARG transfer in the environments. In most cases, a high prevalence of ARB or
520
ARG does not show that gene transfer has happened. In most cases, resistance emergence may
521
be suggested to be due to horizontal gene transfer using ex post facto evidence (Rizzo et al.
522
2013). The major evidence to support ARG transfer include (i) the occurrence of a link between
523
the ARG acquired and mobile genetic elements, including phages, plasmids, transposons
524
(Partridge 2011), (ii) an observation of missing synteny between the DNA that is acquired and
525
the insertion site on the host (Dobrindt et al. 2004; Miriagou et al. 2006; Deurenberg and
526
Stobberingh 2008), (iii) a lack of resemblance between the phylogeny of the resultant host and
527
the supposedly transferred gene (Sørensen et al. 2005; Lal et al. 2008).
528
530
Evidence for the transfer of antibiotic resistance in freshwater environments
531
Direct evidence of transfer is usually scarce, and this is possible because it is difficult to study
532
such a transfer from a donor bacterium to potentially numerous resident bacteria that inhabit
533
complex environments (Sørensen et al. 2005; Rizzo et al. 2013). Generally, several antibiotic
534
resistance determinants that occur in clinical isolates are situated on mobile genetic elements
535
(MGE). This allows their horizontal transfer to other strains (commensals, pathogens, as well as
529
18
536
environmental) or even between bacteria from different taxa. Laboratory microcosm studies
537
are usually applied to demonstrate the transfer of plasmids to susceptible cells at population
538
densities and environmental conditions that are similar to what the bacteria encounter in the
539
natural freshwater environment (Alm et al. 2014). Conjugative plasmids isolated from multi-
540
resistant E.coli strains belonging to the IncP incapability group were transferred to Pseudomonas
541
fluorescens and Aeromonas sp. (Laroche-Ajzenberg et al. 2015). However, microcosm studies
542
show only transfer of a given mobile genetic element (mostly plasmids) and not all the genetic
543
exchanges that can potentially happen in complex microbial communities.
544
Nevertheless, examples of indirect evidence of transfer exists: (i)simultaneous detection of
545
similar resistance genes in pathogens, commensals and environment following the introduction
546
of a given antibiotic in clinics, (ii) increase in the number of bacterial antibiotic resistance genes
547
within the phage genome of the biomes of natural environments, anthropogenic environments
548
and in microbial communities of animal or humans (Muniesa et al. 2013), as well as (iii)
549
concurrent detection of mobile genetic elements positive pathogens in animals (i.e. migratory
550
birds) and their habitat. These are all examples of indirect evidence of transfer (Wu et al. 2018).
551
There are other indirect evidence of the transfer of antibiotic resistance in freshwater
552
environments. In freshwater bodies enteric bacteria are mostly found (Hu et al. 2008; Hoa et al.
553
2011; Suzuki et al. 2013), and they have very high potential to survive over time, long enough
554
for HGT to occur between the enteric bacterial community and aquatic environment (Vital et al.
555
2008). Besides, high species diversity that bacteria that occur in freshwater encourages HGT
556
(Suzuki et al. 2008). Furthermore, in an event of turbulence, faster ARG transfer through HGT
557
can be encouraged by mixing among the bacterial community (Andrup and Andersen 1999).
558 559
Transmission of antibiotic resistant bacteria and resistance genes from the freshwater environment to humans
560
Knowledge of transmission of antibiotic resistant bacteria and resistance genes from the
561
environment to humans is still scarce. A pathogen in a freshwater system may come in contact
562
with a strain that is resistant to antibiotics, for a period long enough for HGT to occur. For
563
freshwater bodies used for recreation, one can acquire an antibiotic resistant pathogen. Once
564
within human system, internal transfer within intestinal microbiota is also possible (Alm et al.
19
565
2014). Laurens et al. (2018) provide an insight into transmission from the environment to
566
humans. They investigated a human case of bacteraemia triggered by IMI-2 carbapenemase-
567
producing Enterobacter asburiae after exposure to river water. The authors applied Pulsed-field
568
gel electrophoresis (PFGE) to compare environmental and clinical bacterial strains. They also
569
used PFGE to determine the blaIMI-2 carbapenemase gene location. Thereafter, they applied
570
fingerprinting technique, 16S rRNA gene PCR–temporal temperature gel electrophoresis to
571
compare the patients' microbiota with those of the bacterial community of the river water to
572
which the patients were exposed. Their result indicated the same plasmidic blaIMI-2 gene was
573
carried in both E.asburiae causing the bacteraemia and that detected in river water a month
574
later. Both river and clinical strains displayed similar PFGE profiles. The patient’s microbiome of
575
carbapenem-resistant bacteria persisted and was autochthonous within the river community (E.
576
asburiae, Pseudomonas fluorescence and Aeromonas veronii). Laurens et al. (2018) then
577
hypothesized that if antibiotic resistance producing strains persisting in various geographic
578
locations are similar to the clinical isolates, then the antibiotic resistant strain may represent a
579
part of the natural reservoir of the resistance, and could be a vehicle for transferring resistance
580
between aquatic environment and human or animal. In addition, they argued that the flanking
581
of the blaIMI-2 gene initially described in E. asburiae by transposable elements on a conjugative
582
plasmid has the potential for the dissemination of gene among bacteria mainly the
583
Enterobacteriaceae (Rotova et al. 2017; Laurens et al. 2018).
584
With regard to ARBs, they can contaminate drinking water treatment plants and distribution
585
systems through freshwater. Drinking water treatment facilities source their raw water from
586
surface water including water from rivers and dams. The activities in the catchment area, as
587
well as the concomitant run-off, determine the level of pollution in surface water. Antibiotic
588
resistant pathogens can enter surface water from human waste (e.g. septic tank and sewage),
589
animal waste (e.g. animal dropping) and intensive farming practices (e.g. dairying and feedlots).
590
Insufficient treatment of water, including a failure to manage turbidity, and inadequate
591
chlorination can lead to pathogens being distributed through municipal water supply (Schwartz
592
et al. 2003; Nadiabartholomew et al. 2014). Schwartz et al. (2003) investigated wastewater,
593
surface water, and drinking water within one municipal system for the presence of resistant
594
bacteria and resistance genes. The authors detected vanA resistance gene without the
595
corresponding detection of bacteria (enterococci) in biofilms sampled from WWTP, wastewater 20
596
effluent, the effluent receiving river and the drinking water supply using the river as source
597
water. The result of the study by Schwartz et al. (2003) thus indicates a possible gene transfer to
598
indigenous drinking water bacteria.
599 600
Mitigation strategies for freshwater environment antibiotic resistance development
601
Antibiotic resistance spreads rapidly, such that even the smallest use of antibiotics significantly
602
increases the development and spread of antibiotic resistance. Limiting antibiotic use and
603
restricting ARG dissemination are the two main well-known methods to mitigate resistance
604
dissemination (Vikesland et al. 2017). Wastewater treatment plants (WWTPs) and hospital
605
effluents are globally accepted as major contributors of ARBs and ARGs in receiving aquatic
606
environments (Czekalski et al. 2012, 2014, 2015; Devarajan et al. 2016). More so, livestock
607
farming and other discharges from urban activities are pollutant sources of ARGs in freshwater
608
bodies, impacting on the resistance of microorganisms within the environment (Amy Pruden et
609
al. 2006; Zhang et al. 2013). Therefore, intervention could target these sources of ARGs in the
610
environment. Strategies to mitigate freshwater antibiotic resistance development could include
611
both traditional and inventive public health approaches, such as secondary treatment of
612
wastewater, providing and implementing standards and guidelines relating to discharges and
613
effluents quality and simple hygiene practices that eradicate or lessen the risk of contamination.
614
Creating methods that are more efficient for the treatment of wastewater from domestic,
615
hospital and industries that contain antimicrobial agents and ARGs before discharge into
616
freshwater is necessary. Disinfection processes that are used to eliminate pathogens during
617
wastewater treatment, including ozonation or chlorination are inadequate in destroying all the
618
genetic material in wastewater under most of the present conditions of operation (Xi et al.
619
2009; Dodd 2012; McKinney and Pruden 2012; Li et al. 2016; Pak et al. 2016; Chang et al. 2017).
620
More so, ozonation and chlorination can select for antibiotic resistance. In order to restrict ARGs
621
dissemination, additional extensive degradation of ARGs in wastewater is critical to address
622
antibiotic resistance. If traditional disinfection processes of wastewater treatment is
623
unavoidable, it is critical that bacteria from the system be eliminated before discharge (Pruden
624
et al. 2013).
21
625
Reducing antibiotics usage in livestock farming will decrease selection pressures and
626
consequently will reduce the preservation of ARGS within the host so as to attenuate resistant
627
strains over time (Bengtsson-Palme and Larsson 2016; Vikesland et al. 2017).
628
antibiotics usage in livestock farming can be achieved by sustaining good animal health and
629
reducing the incidence of disease through the improvement of the conditions under which the
630
animals are bred, and the use of substitutes to antibiotics (Thanner et al. 2016; Tullo et al.
631
2019). Additionally, distancing livestock farming activity from freshwater systems is another
632
preventive measure of antibiotic resistance spread.
633
At a policy level, in order to curtail the spread of antimicrobial resistance from these sources
634
standards on the concentration of antibiotics, which are selecting agents for ARBs, in
635
wastewater treatment plants and hospital effluents to be discharged into nearby aquatic
636
environments must be established. It may be valuable to enforce more restrictions on the use of
637
those antibiotics that persist in the environment, such as fluoroquinolones. Pharmaceutical
638
industries must practice good manufacturing practices that include consideration of the
639
environment, as this could be of benefit (Pruden et al. 2013). So far, it is understood that
640
endorsing regulations on management of effluents from WWTP and even pharmaceutical
641
industries is still a challenge because the cooperation, agreement and enforcement by a large
642
number of stakeholders will be required (Pruden et al. 2013).
643
Conclusion
644
Antibiotics resistant bacteria and resistance genes are prevalent in freshwater environments,
645
and anticipative natural processes of degradation and dilution are not able to completely
646
eradicate them. ARBs are resistant to inactivation through microbial and sunlight degradation,
647
and even if they are effectively inactivated by sunlight, their ARG-containing DNA can still be
648
intact and capable of transferring resistance to non-resistant strains. Rivers with high water
649
quantity can dilute ARGs, but the assimilating capacity of the freshwater can only cope with a
650
certain quantity of pollutants. The ARBs found in freshwater environments might constitute part
651
of the natural reservoir of antibiotic resistance and can act as a vehicle between freshwater
652
bodies and human microbiota. The evidence provided in this study sustains that direct
653
transmission of indigenous freshwater ARBs to humans as well as their transitory insertion in the
654
microbiota can occur. This study provides a baseline knowledge that is fundamental for any risk 22
Reducing
655
assessment study of freshwater bodies with regard to ARGs and ARB. It would be prime if
656
antibiotic usage is reduced to mitigate resistance dissemination.
657
hundreds of millions of people lack access to treated water that is safe for consumption and so
658
relies on freshwater resources for drinking, sustaining crops through irrigation and providing
659
food in the form of fish. The consumption of water that is polluted with ARBs and ARGs or food
660
irrigated with the contaminated water may facilitate the dissemination of antibiotic resistance
661
to humans.
662
References
663
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44
Table 1 Clinically relevant ARBs detected in freshwater bodies ARB group
Extended-spectrum βlactamase producing Enterobacteriaceae
ARG carried
Specie
No
No
of isolates
positive
E. coli
9
6
K. pneumoniae
1
1
E. cloacae
1
1
E. coli
38
38
K. pneumoniae
2
2
E. cloacae
1
1
E. coli
5
1
C. frendii
2
1
E. coli
5
2
K. pneumoniae
2
1
blaTEM−1
K. pneumoniae
2
1
blaCMY−2
E. coli
5
1
blaCTX-M-15
blaCTX-M-14-like
blaSHV
blaSHV−1 in association with blaCTX−M
1
Type freshwater
of
References
Stream
Miyagi and Hirai (2019)
River water
Ye et al. (2017)
C. frendii
2
1
E. coli
5
1
blaSHV
E. coli
74
3
blaCTX−M
E. coli
74
71
blaIMI‐2
Enterobacter asburiae
7
Almost all of the isolates in this study had at least one of the genes (blaTEM, blaCTX-M, blaSHV and blaOxA)
E. coli
167
Klebsiella pneumoniae
114
Citrobacter freundii,
9
Enterobacter cloacae
6
blaSHV−1, blaTEM−1, simultaneously
2
and
blaCTX−M−65
River
Zurfluh et al. (2013)
7
River
Harmon et al. (2018)
Not applicable here
River
Chen et (2010)
al.
Citrobacter koseri
4
Salmonella ssp. Arizonae
3
choleraesuis
Serratia liquefaciens
3
Pantoea spp.
1
blaCTX-M-14
E. coli
30
10
River
Dhanji (2011)
et
al.
blaVEB
Aeromonas spp
29
11
River
Girlich (2011)
et
al.
blaSHV-12
29
10
blaPER-1
29
3
blaTLA-2
29
1
blaGES-7
29
1
Not applicable
Not applicable
River
Tacão et (2012)
al.
blaCTX-M, blaTEM, blaOXA-21
A. hydrophila
integrase genes intI1
Escherichia coli Pseudomonas sp Acinetobacter sp
3
Carbapenemase-producing Enterobacteriaceae
methicillin resistant Staphylococcus aureus
blaKPC-2
E. coli
3
3
K. pneumoniae
1
1
Enterobacter cloacae
1
1
blaIMI-2
Enterobacter cloacae
1
1
blaKPC-2 and blaVIM-1
Klebsiella oxytoca
2
2
mecA gene
Staphylococcus aureus
22
mecA
Staphylococcus aureus
mecA
Staphylococcus aureus
River sediments
Piedra-Carrasco et al. (2017)
22
freshwater beach
Levin-Edens et al. (2012)
698
12
River
Hatcher et al. (2016)
70
24
Freshwater recreational beach
Thapaliya et al. (2017)
70
15
32
27
River
Novais (2005)
vanB
32
4
vanC1
32
1
PVL
Vancomycin Enterococci
resistant
vanA
Enterococci faecalis
4
et
al.
vanB
E. faecalis
Not applicable
Not applicable
vanC
333
166
Van C1
333
10
vanC2/C3
E. casseliflavus
333
164
VanC
Enterococcus spp
216
61
216
1
Van C1
Lata et (2009)
al.
River
Nishiyama et al. (2017)
River
Nam et (2013)
al.
al.
vanA
Enterococcus faecium
20
1
River
Morris (2012)
vanC1
Enterococcus casseliflavus/gallinarum
Not applicable
Not applicable
River
Roberts et al. (2009)
vanC2/3
Enterococcus casseliflavus/gallinarum
5
et
49.7
River
40.3
50.9
39.5
32.8
16S
rRNA Devarajan
gene
sediments
normalized log copy
6
et al. (2016)
Table 1Reported prevalence of antibiotic resistant bacteria in freshwater ecosystems Type of freshwater
Antibiotics
Pond
Carbapenem
% Antibiotic resistant bacteria
References
0.04 Harmon et al. (2019)
0.10 Pond
0.01 Pond
0.05 Dam
0.00 Spring
0.00 Lake
3.24 Lake
0.01 Lake
0.09 Lake
0.02 Natural Pool
0.45 River
Tacão et al. (2015)
77.27 Lake
Penicillin
Pang et al. (2015)
64.55 Ampicillin
34.55 Cephalothin
14.55 Chloramphenicol
77.27 Tetracycline Rifampicin
1.00
1
Type freshwater
of
Antibiotics
% Antibiotic resistant bacteria
References
6.6–21.0 River
Ampicillin
Ash et al. (2002) and Aubron et al. (2005)
10.4–25.7 22.6–24.1 34.5–38.4 26.9 59.2 26.4 13.7–34 10.1–36.6 4.9–52.5 5.9 19.7–23.7 6.7–73.0 6.1–21.5 20.0–53.0 12.4–20.0
2
Type of freshwater
Antibiotics
%
References
Antibiotic resistant bacteria
15.7 River
Ampicillin
Ash et al. (2002) and Aubron et al. (2005) 32 3.5–33.9 3.9 17.0–25.0
22.5
93 Cephalothin 96 94 91 8 86 25 73 67 77
3
Type freshwater
of
Antibiotics
% Antibiotic resistant bacteria
References
78 River
Cephalothin
Ash et al. (2002) and Aubron et al. (2005)
36 96 86 38 98 94 96 5 Cefotaxime
6 13 9 25 12 0 5
4
Type of freshwater
Antibiotics
%
References
Antibiotic resistant bacteria
9 River
Cefotaxime
Ash et al. (2002) and Aubron et al. (2005) 20 21 4 0 8 0 16 2 0 4 7 39 2
Ceftazidime 0
5
Type of freshwater
Antibiotics
River
Ceftazidime
% Antibiotic resistant bacteria
References
1 Ash et al. (2002) and Aubron et al. (2005) 6 0 0 8 5 2 0 9 0 2 0 0 0 0 0
6
Type of freshwater
Antibiotics
%
References
Antibiotic resistant bacteria
5 River
imipenem
Ash et al. (2002) and Aubron et al. (2005) 0 0 10 0 0 4 1 0 0 6
68 Amoxicillin+clavalanic 8 acid 10 37 36
7
Type of freshwater
Antibiotics
% Antibiotic resistant bacteria
River
0 Amoxicillin+clavalanic 24 acid
References
Ash et al. (2002) and Aubron et al. (2005)
8
Table 1 Abundance of antibiotic resistance genes detected in freshwater bodies (sul, tet, aa, bla, flo as well as erm and mef signify sulfonamide, tetracycline, aminoglycoside, β-lactam, chloramphenicol as well as Macrolide-Lincosamide-Streptogramin) Type of fresh-water body River sediments
Type of ARG detected sul1
Quantity detected
Unit of measurement
10− 6 - 10− 3
Gene copy numbers normalized to bacterial 16S rRNA gene copy number
References
Amy Pruden et al. (2006); Pei
et
al.
(2006);
Storteboom et al. (2010)
Lake sediments
blaSHV blaTEM blaCTX-M blaSHV blaNDM aadA blaSHV blaTEM blaCTX-M blaSHV blaNDM aadA sul1
0–6.4 4.7–6.1 1.7–6.1 0–6.4 3.5–5.1 2.2–5.2 39.5 49.7 50.9 39.5 32.8 40.3 7.3 × 10− 1 ± 9 × 10− 2
sul2
4 × 10− 2 ± 3 × 10
sul1 tet(B) blaSHV
2.2 × 109 1.5 × 106 1.4 × 10–6 to 8.9 × 10–4
blaTEM blaCTX-M blaSHV blaNDM aadA
2.6 × 10-5 to 2.8 × 10–3 3.2 × 10-6 to 1.7 × 10–3 1.4 × 10–6 to 8.9 × 10–4 6.07 × 10–6 and 1.2 × 10–6 from 1.5 × 10–5 to 1.1 × 10–3
16S rRNA gene normalized log copy
Devarajan et al. (2016)
16S rRNA gene normalized log copy
Devarajan et al. (2016)
Gene copy number normalized to bacterial 16S rRNA gene copy numbers
Czekalski et al. (2012)
ARG copies g−1 wwt
Czekalski et al. (2014)
16S rRNA gene normalized log copy
Devarajan et al. (2015)
Table 2 continued. Abundance of antibiotic resistance genes detected in freshwater bodies (sul, tet, aa, bla, flo as well as erm and mef signify sulfonamide, tetracycline, aminoglycoside, β-lactam, chloramphenicol as well as Macrolide-Lincosamide-Streptogramin)
Type of fresh-water body Lake water
Sediments and river water
Type of ARG detected sul1
Quantity detected
Unit of measurement
References
1.5 × 10−3 - 2.1 × 10
Gene copy numbers normalized to bacterial 16S rRNA gene copy numbers
Czekalski et al. (2015)
sul2 blaTEM
Up to 3.4 × 10−3 Both samples analysed were positive for these ARGs.
log10 GC/mL or g) of each ARG in the bacterial (BAC)
Piedra-Carrasco et al. (2017)
qnrA qnrS mecA blaCTX-M blaSHV blaNDM Groundwater sul1 sul2 tet(A) tet(C) tet(O) tet(W) blaSHV floR ermA mefA
The abundances of copies/16S rRNA gene these ranged from copies 6.61 × 10−7 to 2.30 × 10−1
Szekeres et al. (2018)
Figure 1 A depiction of the potential origin, spread, transmission and degradation/dilution of ARBs and ARGs
Freshwater environment is a reservoir where antibiotics impact microorganisms Carbapenem resistant bacteria are less prevalent in freshwater environments Different ARGs, sul1, sul2, tet(B), aadA, blaTEM, blaCTX-M, blaSHV, blaNDM are detectable in Freshwater. Natural processes of degradation and dilution not able to eradicate ARBs and ARGs in freshwater. The ARBs can act as a vehicle between freshwater bodies and human microbiota. The major contributors of ARBs and ARGs are points for mitigating antibiotic resistance development.
This study provides knowledge that is fundamental for any risk assessment study estimating the actual risk of antibiotic resistance dissemination from freshwater reservoirs.
The authors declare that there is no conflict of interest.
S0269-7491(19)31443-5
DOI:
https://doi.org/10.1016/j.envpol.2019.113067
Reference:
ENPO 113067
To appear in:
Environmental Pollution
Received Date: 19 March 2019 Revised Date:
13 August 2019
Accepted Date: 15 August 2019
Please cite this article as: Nnadozie, C.F., Odume, O.N., Freshwater environments as reservoirs of antibiotic resistant bacteria and their role in the dissemination of antibiotic resistance genes, Environmental Pollution (2019), doi: https://doi.org/10.1016/j.envpol.2019.113067. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.
Sunlight degradation fate dilution Freshwater pool of ARB and ARGs
1 2 3
Freshwater environments as reservoirs of antibiotic resistant bacteria and their role in the dissemination of antibiotic resistance genes Chika F. Nnadozie* and Oghenekaro Nelson Odume
4 5 6
Unilever Centre for Environmental Water Quality, Institute for Water Research, Rhodes University, PO Box 94, Grahamstown 6140, South Africa Corresponding author*: E-mail: [email protected]; Tel.: +27 738052544
7 8
Abstract
9
Freshwater environments are susceptible to possible contamination by residual antibiotics that
10
are released through different sources, such as agricultural runoffs, sewage discharges and
11
leaching from nearby farms. Freshwater environment can thus become reservoirs where an
12
antibiotic impact microorganisms, and is an important public health concern. Degradation and
13
dilution processes fundamental for predicting the actual risk of antibiotic resistance
14
dissemination from freshwater reservoirs. This study reviews major approaches for detecting
15
and quantifying antibiotic resistance bacteria (ARBs) and genes (ARGs) in freshwater and their
16
prevalence in these environments. Finally, the role of dilution, degradation, transmission and
17
the persistence and fate of ARB/ARG in these environments are also reviewed. Culture-based
18
single strain approaches and molecular techniques that include polymerase chain reaction
19
(PCR), quantitative polymerase chain reaction (qPCR) and metagenomics are techniques for
20
quantifying ARB and ARGs in freshwater environments. The level of ARBs is extremely high in
21
most of the river systems (up to 98% of the total detected bacteria), followed by lakes (up to
22
77% of the total detected bacteria), compared to dam, pond, and spring (< 1%). Of most concern
23
is the occurrence of extended-spectrum β-lactamase producing Enterobacteriaceae, methicillin
24
resistant Staphylococcus aureus (MRSA) and vancomycin resistant Enterococcus (VRE), which
25
cause highly epidemic infections. Dilution and natural degradation do not completely eradicate
26
ARBs and ARGs in the freshwater environment. Even if the ARBs in freshwater are effectively
27
inactivated by sunlight, their ARG-containing DNA can still be intact and capable of transferring
28
resistance to non-resistant strains. Antibiotic resistance persists and is preserved in freshwater
29
bodies polluted with high concentrations of antibiotics. Direct transmission of indigenous
30
freshwater ARBs to humans as well as their transitory insertion in the microbiota can occur.
31
These findings are disturbing especially for people that rely on freshwater resources for
32
drinking, crop irrigation, and food in form of fish. 1
33 34
Introduction
35
The continual overuse of antibiotics is causing their release into the environment. The release of
36
antibiotics into the environment is a concern because of the consequent development of
37
antibiotic resistance genes and bacteria that lead to the reduced effect of antibiotics on human
38
and animal pathogens. The prevalence of antibiotics in the environment is of growing global
39
concern both for the public and research community. Freshwater environments are among the
40
natural environments that are susceptible to possible contamination with antibiotics released
41
through different sources, such as agricultural runoffs, sewage discharges and leaching from
42
nearby farms. The presence of antibiotics in the environment, a higher density of active bacteria
43
community indigenous to freshwater, creates an environment that is suitable for the
44
development of antibiotic resistance genes (Pereira et al. 2013; Marti et al. 2014; Chen et al.
45
2017). Freshwater ecosystems have thus become hotspots for horizontal gene transfer (HGT) of
46
antibiotic resistance genes, and consequently where resistance evolution occurs. The potential
47
for ARBs and ARGs to persist in freshwater environments can lead to increased risk of infections
48
with resistant pathogens. ARGs can persist in the environment and ultimately return to human
49
as well as animal pathogens. The same freshwater body that is receiving wastewater also serves
50
as a source of a drinking water reservoir and even recreation (Baquero et al. 2008). In addition,
51
the spread of ARB and ARG can result in high ARG pool in environmental bacteria, therefore
52
encouraging the transfer of resistance into well-known as well as emerging pathogens (Czekalski
53
et al. 2012).
54
Investigations on the occurrences and type of antibiotic resistance in freshwater environments
55
is limited compared to clinical settings, such as common indicator organisms (Enterococci and
56
coliforms), bacterial pathogens that cause infections that are difficult to treat in humans (Araújo
57
et al. 2010; Adelowo et al. 2014; Ndlovu et al. 2015; Berendonk et al. 2015; Chen et al. 2017).
58
The limited investigation of antibiotics in freshwater environments could be because antibiotic
59
concentrations in these environments are presumed to be naturally low in Freshwater
60
environments. However, even if the antibiotic concentration in freshwater environments is low,
61
the selection of resistant bacteria can occur (Kümmerer 2009; Marti et al. 2014). The majority
62
of studies on the occurrence and fate of ARBs and ARGs in freshwater environments, including 2
63
streams, lakes and rivers, are focussed mostly on discharge point of wastewater treatment
64
plants (WWTPs) (LaPara et al. 2011; Marti et al. 2013; Chen et al. 2013; Sabri et al. 2018).
65
Concentrations of ARBs and ARGs are usually high at the point of effluent discharges from
66
WWTWP into streams, but gradually become reduced downstream of the discharge point
67
(LaPara et al. 2011; Marti et al. 2013; Chen et al. 2013; Sabri et al. 2018). The decline in
68
concentration of ARGs and ARBs downstream of WWTW effluent discharge points can be as a
69
result of different factors such as dilution, degradation, adsorption, transport (Floehr et al. 2013;
70
Anyaduba 2016; Nelson et al. 2018; Yoon et al. 2018). Among these different factors, dilution
71
and degradation play a significant role in the fate of ARGs in freshwater environments (Snow et
72
al. 2015; LaPara et al. 2015; Anyaduba 2016; Jerde et al. 2016; Collins et al. 2018) (Figure 1).
73
Thus, this paper investigates how dilution and degradation influence the fate of ARB, and ARG,
74
and the subsequent development and dissemination of resistance to human pathogens. Before
75
proceeding to evaluate the influences of dilution and degradation on ARB and ARG, we first
76
present an analytic review of techniques for detecting and quantifying ARBs and ARGs in
77
freshwater environments. The abundance and diversity of resistance genes in freshwater
78
environments are also reviewed.
79 80
Techniques for quantifying antibiotic resistance genes and bacteria in freshwater bodies
81
Culture-based single strain approaches are the most frequently applied techniques for
82
quantifying ARB and ARGs in freshwater environments (West et al. 2010; Suzuki et al. 2013;
83
Zurfluh et al. 2013). Culture-based approaches involve growing the microorganisms on a
84
nutrient medium (most preferably a growth medium that is selective to the bacteria of interest).
85
Individual isolates are tested for antibiotic resistance by growing them on nutrient medium
86
supplemented with the antibiotics of interest (West et al. 2010), or embedded within an
87
antibiotic disk (Narciso-da-Rocha and Manaia 2016). In most cases, the antibiotics are those
88
commonly used in human as well as veterinary medicine that show different mechanisms and
89
pathways for activity, such as inhibition of cell wall synthesis or protein synthesis (Moore 2013).
90
The culture-based approach is limited because it does not capture the major components of the
91
natural assemblage of a freshwater microbial community, and ARGs detected through such a
92
method is just a selection of the total bacteria community (Suzuki et al. 2013). Alternative
3
93
approaches involve the application of molecular tools , including Polymerase chain reaction and
94
quantitative polymerase chain reaction (PCR and qPCR) (Marti and Balcázar 2013; Suzuki et al.
95
2013; Mao et al. 2014; Czekalski et al. 2015; Xiong et al. 2015; Di Cesare et al. 2017; Zhou et al.
96
2017; Giebułtowicz et al. 2018; Sabri et al. 2018; Zheng et al. 2018; Liu et al. 2019), and
97
metagenomics profiling ARB and ARG of the environment (Kristiansson et al. 2011; Port et al.
98
2012; Amos et al. 2014, 2018; Bengtsson-Palme et al. 2014; Xiong et al. 2015; Fitzpatrick and
99
Walsh 2016; Zheng et al. 2017; Fresia et al. 2018).
100
Polymerase chain reaction (PCR) uses small quantity of a sample DNA (template), two primers
101
flanking the target sequence (ARG and ARB), nucleotides, together with thermostable DNA
102
polymerase to amplify a particular region of DNA of interest, creating a large quantity of DNA
103
from a very small environmental sample (Boyle 2014). qPCR is an improvement of the
104
conventional PCR ,of which the targeted gene is amplified and quantified at the same time
105
(Pabinger et al. 2014). Quantification of the number of the target present in template DNA is
106
extremely challenging with conventional PCR. With qPCR, the quantity of the product that is
107
produced is observed in the course of the reaction by observing the fluorescence of dyes or
108
probes that are added in the reaction mix. The amount of fluorescence of dyes or probes is
109
proportional to the amount of product formed, and the amount of amplification cycles
110
necessary for obtaining a particular number of DNA molecules is recorded. By assuming a
111
certain amplification efficiency, one can calculate the amount of DNA molecule of the amplified
112
target sequence that is originally present in the sample (Kubista et al. 2006). The lesser the PCR
113
cycles that is needed to make enough material for detection, the more the copies of DNA
114
template that is present at the beginning of the experiments (Pabinger et al. 2014)
115
Metagenomics involves characterizing the total microbial community DNA (which includes
116
prokaryotes, viruses, and eukaryotes) in an environment (Handelsman et al. 1998; Staley and
117
Sadowsky 2016). Thus, using a single dataset of DNA sequence, one can quantify all known
118
resistance genes present in the sample (Munk et al. 2017). Metagenomics can be through
119
sequence-based or function based screening. A sequence-based metagenomic uses sequencing-
120
by-synthesis technology to produce millions of sequence reads without the need for cloning
121
(Gilbert and Dupont 2011). Screening through this approach does not require prior knowledge
122
of the sequence of interest. Alternatively, a function-based screening is a method whereby DNA 4
123
clones of the environmental library are screened for their ability to confer a function of interest
124
to a heterologous host. The advantage of this approach is that it does not require prior
125
knowledge of the sequence. Therefore, it provides direct evidence of the function of genes that
126
are not previously known (Martínez and Osburne 2013). Functional screening, whereby fosmid
127
libraries were constructed and screened to evaluate the level of antibiotic resistance have been
128
undertaken for river water and sediments (Amos et al. 2014; González-Plaza et al. 2018).
129
Compared to a sequence-based metagenomic screening, a function-based screening requires
130
less amount of DNA from environmental samples (Staley et al. 2015).
131
Metagenomics offers advantages over PCR because it is possible to quantify thousands of target
132
genes in environmental samples without prior knowledge of bacteria or genes present in the
133
sample. However, this approach is limited in several ways: i) determining the function of newly
134
discovered sequences can be a challenge, ii) large fractions of the metagenomic sequences
135
generated usually do not show any significant similarity to previously annotated sequences, thus
136
making annotation difficult (Nnadozie et al. 2017).
137 138
Prevalence of antibiotic resistant bacteria in freshwater ecosystems
139
Freshwater environments are an important reservoir of ARBs and ARGs. Table 1 reflects the
140
prevalence of ARBs in freshwater ecosystems. One can observe that they are widespread in
141
different types of freshwater environments. The level of ARBs is extremely high in most of the
142
river systems (up to 98% of the total detected bacteria), followed by lakes (up to 77% of the
143
total detected bacteria). This indicates that rivers and lakes serve as a significant reservoir for
144
the spread of antibiotic resistance to opportunistic pathogens. The high concentration of ARBs
145
that are resistant to cephalothin, penicillin, tetracycline, ampicillin, and chloramphenicol in
146
rivers and lakes, indicate that these antibiotics are present in the freshwater systems. This could
147
be because of the widespread use of these antibiotics as basic antimicrobial drugs. The
148
widespread application of antibiotics leads to a high proportion of bacteria are resistant to
149
them. A number of factors including the geographical location of the region and prescription
150
policies can influence the proportion of ARBs in freshwater environments. In regions where
151
restrictive rules exist regarding prescription and disposal, a low prevalence of ARB has been
152
reported (Tacão et al. 2015). 5
153
On the other hand, the concentration of ARBs in the dam, pond, and spring (< 1%) are relatively
154
low (Table1). Yet, dam and pond systems could be underappreciated reservoir s of ARBs. The
155
prevalence of ARBs could be high in these environments, considering that usually only a sample
156
is withdrawn during sampling. The reason for the higher frequency of ARBs in river and lake
157
compared to the dam and pond require further investigation.
158
Of most concern to public health is the occurrence of extended-spectrum β-lactamase
159
producing Enterobacteriaceae, Carbapenemase-producing Enterobacteriaceae, methicillin
160
resistant Staphylococcus aureus (MRSA) and vancomycin resistant Enterococcus (VRE) – all of
161
which can cause potentially highly epidemic infections (Arora et al. 2014) (Table 2). A worldwide
162
environmental dissemination of these ARBs hosting ARGs has been suggested in recent studies
163
(Spindler et al. 2012; Zhang et al. 2013; Zurfluh et al. 2013; Khan et al. 2013; Czekalski et al.
164
2014; Laht et al. 2014; Devarajan et al. 2015, 2016; Djenadi 2017). Their presence in freshwater
165
environments has been reported in countries around the world, such as Portugal, Finland,
166
France, Pakistan, Switzerland, United States of America, China and Brazil (Spindler et al. 2012;
167
Zhang et al. 2013; Zurfluh et al. 2013; Khan et al. 2013; Czekalski et al. 2014; Laht et al. 2014;
168
Devarajan et al. 2015, 2016; Djenadi 2017).
169
The family Enterobacteriaceae are Gram-negative bacteria, including Escherichia coli (E. coli),
170
Citrobacter, Salmonella, Yersinia pestis, Shigella, Proteus, Enterobacter, Salmonella and
171
Klebsiella (Patel and Nagel 2015). Extended-spectrum β-lactamase and Carbapenemase-
172
producing Enterobacteriaceae (CPE), such as E. coli, K. pneumoniae, E. cloacae, C. frendii,
173
Enterobacter asburiae and Klebsiella oxytoca isolates have been detected in freshwater
174
environments such as streams and rivers (Table 2) (Zurfluh et al. 2013; Ye et al. 2017; Harmon et
175
al. 2019). The β-lactamase-producing Enterobacteriaceae are resistant to the modern extended-
176
spectrum cephalosporins antibiotics as well as penicillins and monobactams (Bush and Jacoby
177
2010; Miyagi and Hirai 2019). Their resistance to these antibiotics is due to their plasmid-
178
mediated production of enzymes that hydrolyze the β-lactam ring of the antibiotic compounds.
179
This type of resistance is caused by a high number of point mutation variations of established
180
broad spectrum β –lactamases – the so-called extended-spectrum β-lactamases (ESBLs). Several
181
ESBLs are members of SHV and TEM β-lactamases, while others are classified as OXA, CTX-M,
182
GES, TLA and PER β-lactamases (Poirel et al. 2000; Coque et al. 2008; Bush and Jacoby 2010). 6
183
The reported high occurrence of ESBL in Zurfluh et al. (2013) is disturbing because the study was
184
carried out in Switzerland where strict policies apply to antibiotic use (Filippini et al. 2006;
185
Zurfluh et al. 2013). The identification of ESBL genes in environmental Aeromonas,
186
Pseudomonas sp., and Acinetobacter isolates revealed an extensive assortment of mobilization
187
events, implying that these bacteria are vehicles for ESBL dissemination (Girlich et al. 2011;
188
Tacão et al. 2012). Although the number of studies exploring freshwater as possible reservoirs of
189
ESBL producers is still scarce, their reported presence in the freshwater environments is highly
190
disturbing because of the potential to transmit their genes into human pathogens (Flores
191
Ribeiro et al. 2012; Zhou et al. 2014; Czekalski et al. 2015; Chen et al. 2015, 2017).
192
Furthermore, carbapenemases are a large group of β-lactamases that are grouped into classes
193
ABC and D that inactivate carbapenem antibiotics. Carbapenem antibiotics, such as imipenem
194
and meropenem are used as last resort drugs to treat dangerous infections caused by ESBL
195
producers. Due to the limited antibiotic options, the treatment of infections caused by the
196
organisms that are resistant to carbapenem present serious challenges (Izadpanah and Khalili
197
2015). CPE has been designated as urgent threats (Centers for Disease Control and Prevention
198
2013, 2018) and is associated with very high mortality rates. Therefore the presence of CPE in
199
the freshwater bodies presents a huge challenge to public health (Harmon et al. 2019). Table 2
200
shows that the CPE including E. coli, K. pneumoniae, K. oxytoca and E. cloacae that are positive
201
for KPC-2, VIM- and IMI-2 were recovered in the bottom sediments of freshwater environments.
202
Increased loading and persistence have been implicated for their recoveries in the sediments
203
(Piedra-Carrasco et al. 2017). The IMI and KPC enzymes belong to Class A, which are inhibited by
204
clavulanic acid. The KPC enzymes are plasmid coded. The VIM types belong to Class B
205
carbapenemases and are integron coded (Bush and Jacoby 2010; Zurfluh et al. 2013). The
206
increased loading and persistence of CPE implies that these ARBs are detectable in sediments,
207
even if they are not found in water columns. Klebsiella and Enterobacter seemed to be the most
208
prevalent microorganism that produces KPC-2 (Picão et al. 2013)
209
The occurrence of methicillin-resistant Staphylococcus aureus (MRSA) in the recreational
210
freshwater environment suggests a potential colonization of people that come in contact with
211
them, and possible environmental contamination (Levin-Edens et al. 2012; Fogarty et al. 2015;
212
Hatcher et al. 2016; Thapaliya et al. 2017). Infections caused by MRSA are difficult to treat. 7
213
Staphylococcus strains that carry methicillin resistance gene mecA on their staphylococcal
214
cassette chromosome of type II cause healthcare-associated MRSA infections. They are resistant
215
to many classes of antibiotics such as oxacillin methicillin, penicillin, and amoxicillin.
216
Alternatively, Staphylococcus strains that carry Panton-Valentine leukocidin (PVL) genes on their
217
SCCmec type IV cause community-associated MRSA (CA-MRSA) infections, and are resistant to
218
fewer classes of antibiotics.
219
In addition, antimicrobial–resistant Enterococci spp. isolates have been recovered in freshwater
220
environments (Table 2). For example, vanCtype VRE is widely distributed in aquatic
221
environments, including rivers and coastal areas (Zdragas et al. 2008; Nam et al. 2013;
222
Nishiyama et al. 2017). Vancomycin used to be the most effective last-line-of-defense
223
antimicrobials for managing antimicrobial resistant Enterococci, but its efficiency is now in doubt
224
(Arias and Murray 2012). vanA, vanB, vanC, vanD, vanE and vanG are the six recognised
225
vancomycin-resistant genes (Khan et al. 2008). vanA genes are associated with non-conjugative
226
and conjugative plasmids carrying Tn1546‐like transposons, whereas vanB genes are associated
227
with conjugative transposons, which includes Tn5382, Tn1549 and Tn1547 (Roberts et al. 2009).
228
At this juncture, it is noteworthy that the levels of ARGs presented in this section are based on
229
either total heterotrophic bacteria counts or particular pathogenic bacteria cultured in the
230
laboratory and must be viewed as representing a fraction of bacteria that actually occur in
231
freshwater environments.
232 233
The abundance of antibiotic resistance genes in freshwater environments detected using non-culture based techniques
234
Using
235
chloramphenicol as well as Macrolide-Lincosamide-Streptogramin resistance genes have been
236
widely reported in freshwater bodies (Amy Pruden et al. 2006; Pei et al. 2006; Storteboom et al.
237
2010; Czekalski et al. 2012, 2014, 2015; Szekeres et al. 2018) (Table 3). The majority of studies
238
have focused on sulfonamide resistance genes (sul1 and sul2) compared to the other resistance
239
genes. The extensive focus on sulfonamide resistance genes could be attributed to the fact that
240
they are considered as important indicators of freshwater pollution. Sulfonamide resistance
241
genes are less abundant in pristine environments. sul1 abundance is usually associated with
242
input from WWTPs effluent discharges into freshwater, while sul2 is mainly associated with
molecular
techniques,
sulfonamides,
8
tetracyclines,
β-lactam,
aminoglycosides,
243
inputs from urban activities such as sand dredging, agricultural runoff, urban discharges,
244
religious rituals, open defecation, and other anthropogenic activities (Devarajan et al. 2016). The
245
presence of sul genes in the freshwater environment is not driven by the commonly known
246
drivers such as nutrient enrichment, pH, and the quality and amount of organic carbon of the
247
compositional bacterial community in the freshwater environment. Czekalski et al. (2015)
248
provide a baseline useful for distinguishing between pollution-induced concentrations of ARG
249
from the natural background levels. Freshwater bodies with sul1 and sul2 abundances of
250
2.8 × 10− 3 ± 3.9 × 10− 4 and 3.1 × 10− 3 ± 3.1 × 10− 3 are considered not impacted by ARG pollution
251
(Czekalski et al. 2015). Thus, sul1 and sul2 abundances ranging 1.5 × 10− 3 to 1.6 × 10− 2 and
252
3.1 × 10− 3–7.2 × 10− 5, respectively are considered baseline. The values provided are ratios of
253
both sul genes normalized to eubacterial 16S rRNA genes.
254
ARGs are considered contaminants of emerging concern because some of the genes are present
255
in bacteria as structurally innate reservoir genes. blaTEM genes are usually obtainable in most
256
environments regardless of whether the site has been exposed to anthropogenic contamination
257
and this has been attributed to their ubiquitous presence as housekeeping genes (Demaneche
258
et al. 2008). aadA and blaTEM genes have also been reported to be present even before the start
259
of the twentieth century (Devarajan et al. 2015). blaCTX-M gene is the most common ESBL
260
globally, with CTX-M-15 and CTX-M-14 mainly invading humans, animals and the environment
261
(Cantón et al. 2012).
262
It is important to note that resistance genes in Table 3 do not necessarily represent the full
263
range of genes freshwater bodies as other ARGs may occur. It is possible that the abundance of
264
other resistance genes is not sufficient enough to be captured by the method applied. The
265
studies in Table 3 applied qPCR methods, which are limited. Reported variations in the
266
abundance of resistance genes could be due to a number of factors including the period of
267
sampling, disposable practices, and sampling season (Devarajan et al. 2016). Studies on factors
268
influencing the variation of abundance of resistance genes in the environment are scarce but
269
necessary for establishing a reliable threshold for analyzing ARG pollution.
9
272
Effects of dilution and degradation on the fate of antibiotic resistant bacteria and antibiotic resistance genes in the freshwater environment
273
The previous section discusses factors that drive the persistence of antibiotic resistance in
274
Freshwater without regarding the issues of degradation, dilution, transportation, and
275
adsorption. These processes are particularly important in making ecologically relevant
276
inferences due to the impacts of water volumes, the presence of substrate surfaces, current and
277
tides (Foote et al. 2012; Jerde et al. 2016; Shogren et al. 2017). Antibiotic resistant genes and
278
antibiotic resistant bacteria in aquatic environments can exist either as intracellularly in viable
279
antibiotic resistant bacteria as genomic and plasmid DNA or extracellularly as free eDNA
280
(environmental DNA) that is shielded within phage capsids, extracellular polymeric substances
281
(EPS), cell debris or on clay mineral surfaces (Dodd 2012). In terms of their fate, the ARBs and
282
ARGs can be transported, experience decay through biological and non-biological processes,
283
adsorbed onto particulate matter, diluted, uptake by aquatic microorganisms through HGT
284
(Chen and Dubnau 2004; LaPara et al. 2015; Anyaduba 2016). Transport means the movement
285
of water by advection or diffusion in streams or ponds, and ARGs and ARBs can be transported
286
over considerable distances in water (Poté et al. 2003). Compared to the other processes,
287
transport does not play a substantial role in the fate of ARBs and ARGS in aquatic environments
288
(Anyaduba 2016). Studies suggest that it is dependent on the flow regime of the water and
289
generally in high flowing rivers the ARBs and ARGs are diluted, and so it is more of dilution effect
290
(LaPara et al. 2015; Jerde et al. 2016). This explains why rivers and lakes located in areas prone
291
to drought are expected to have a greater presence of ARBs and ARGs (Anyaduba 2016).
292
Transport will mainly influence the representativeness of the concentration of ARBs and ARG
293
that is measured at one point of a Freshwater environment (Goldberg et al. 2016). On the other
294
hand, adsorption is the deposition and trapping in substrate crevices, sticking to stream biofilms
295
colonizing substrate surfaces (Anyaduba 2016; Jerde et al. 2016; Shogren et al. 2017). ARGs
296
interact with minerals, humid acid in the freshwater environment, explaining why the majority
297
of ARGs remain immobilized in sediments. Immobilization of ARGs will lead to longer
298
persistence in water (Ficetola et al. 2008). The extended persistence of ARGs by adsorption
299
creates an opportunity for resuspension to occur, for instance when there is shear stress or high
300
flow events (Turner et al. 2015).
270 271
10
301
Degradation and dilution have been found to play major significant role in the control of the fate
302
and persistence of microbial contaminants in Freshwater and are fundamental for predicting the
303
actual risk of antibiotic resistance dissemination from freshwater reservoirs (Foote et al. 2012;
304
Lasagna et al. 2013; Anyaduba 2016; Goldberg et al. 2016). Therefore the following section
305
discusses how degradation and dilution determine the persistence of ARBs and ARGs in
306
freshwater.
307
Degradation
308
Degradation is the most significant process that influences the fate of ARBs and ARGs in the
309
freshwater environment (Goldberg et al. 2016). Degradation can occur due to either microbial
310
activity or natural sunlight exposure.
311
microbiome secretes enzymes during metabolism to break down large organic molecules before
312
they are assimilated into cells. Thus, environmental ARGs are prone to the biotic degradation
313
process that involves DNAses. DNAses catalyze the cleaving of the phosphodiester bond
314
between the phosphate group and the deoxyribose sugars in the DNA, releasing nucleotides
315
that are assimilated by the organisms. Studies have demonstrated DNA degradation in aquatic
316
systems by high microbial and enzymatic activity (Zhu 2006; Dejean et al. 2011; Pilliod et al.
317
2014). ARGs discharged from wastewater treatment facilities into receiving water significantly
318
have a loss of mechanisms due to natural decay (Anyaduba 2016; Jerde et al. 2016; Shogren et
319
al. 2017). Higher degradation rates occur with environmental conditions, such as neutral pH,
320
moderately high temperature and UV-B irradiation concomitantly that favor microbial growth
321
and increase the presence of nucleases, which enhance ARG loss in ecosystems (Strickler et al.
322
2015; Anyaduba 2016). However, sediments provide ARGs protection from biological
323
degradation (Turner et al. 2015). In previous studies, DNA protection from degradation upon
324
adsorption onto the sediments has been explained in different ways. DNA is protected from
325
nucleases on adsorption onto sediments (Torti et al. 2015). It is suggested that adsorption onto
326
sediments reduces accessibility to nucleases (Khanna and Stotzky 1992). Nucleases adsorbed on
327
sediment become inactivated (Lorenz and Wackernagel 1987, 1992; Sarkar et al. 1989;
328
Romanowski et al. 1991; Paget et al. 1992). The DNA can also bind to biogenic sediment
329
components, which include proteins and humic acids that can enhance the resistance to
With regards to microbial activity, organotrophic
11
330
nucleases (Nielsen et al. 2006). Furthermore, the DNA may interact with exopolymeric
331
substances of sediment particles that will shield them from degradation.
332
In the case of natural sunlight, it is a generally accepted disinfectant and has always been relied
333
on as a strong determinant of the persistence of microbial pollutants in surface waters, including
334
fresh and marine surface waters (Boyle et al. 2008; Nelson et al. 2018). Sunlight degradation of
335
bacteria is by photoinactivation and is well studied. Sunlight degradation has been observed in
336
Salmonella enterica, Shigella flexneri, Escherichia coli (E. coli) and Enterococcus faecalis (E.
337
faecalis) after sunlight exposure (Berney et al. 2007; Sassoubre et al. 2014; Mcclary et al. 2017;
338
Scoullos et al. 2019; Busse et al. 2019). More specifically, few studies exist that demonstrate
339
sunlight mediated degradation of bacteria in freshwater.
340
coliforms, E. coli, Legionella spp. Enterococci, F-RNA, and somatic coliphages, phages was
341
observed in freshwater (Dutka 1984; Davies and Evison 1991; Sinton et al. 2002; Dick et al. 2010;
342
Korajkic et al. 2014; Wanjugi et al. 2016).
343
However, sunlight can vary substantially in spectral quality, and underwater. Apart from sunset
344
and sunrise, water particles can change the spectral composition of sunlight. Furthermore, solar
345
radiation differs seasonally (Salter 2018), with longer sunlight duration occurring during summer
346
at high latitudes, and no sunlight at all during winter. Differences in seasonal distribution,
347
strength and biological activity of UV radiation have been recognized in the literature for while
348
(Jablonski and Chaplin 2010). The geometry of sunlight that reaches different places varies in
349
different seasons. An implication of this seasonal differences in solar irradiation underwater is
350
that sunlight mediated degradation of microorganisms in water will vary based on season,
351
weather condition, time of the day and location (Nelson et al. 2018). In the study by Calero-
352
Cáceres et al. (2017), higher prevalence and abundance of ARGs were observed in winter, and
353
this was attributed to lower irradiance and temperature in winter. Besides, a greater total
354
organic carbon in water during winter has been suggested to could cause bacterial regrowth,
355
which leads to a higher abundance of ARGs compared to summer seasons (Calero-Cáceres and
356
Muniesa 2016).
357
Also, different intensities of solar UV radiation are obtainable in different regions. Therefore, in
358
some regions, the intensity of sunlight radiation may not be sufficient to initiate sunlight
12
Sunlight inactivation of faecal
359
mediated degradation. Compared to Eurasia regions, it is suggested that microorganisms in
360
waters located in regions that are near the equator, and in high altitudes experience high UVR
361
(Jablonski and Chaplin 2010; Nguyen et al. 2014). Therefore, more sunlight mediated
362
degradation is expected.
363
Additionally, factors such as cloudiness, aerosols, stratospheric ozone (Zepp et al. 2018), water
364
quality, snow and ice over, and depth at which the microorganisms reflexively circulate can
365
modify the exposure of the microorganisms to solar radiation (Sulzberger et al. 2019;
366
Williamson et al. 2019). Snow-cover on ice can inhibit most or all UV radiation from entering the
367
water column, depending on how thick it is. Regarding mixed layer depth, the exposure of
368
microorganisms to sunlight mediated degradation depends on the vertical position within the
369
water column. Microorganisms that are trapped near the surface are more exposed to sunlight
370
mediated degradation (Williamson et al. 2019).
371
Furthermore, the water quality and depth exert appreciable influence over the rate of
372
photoinactivation. The presence of dissolved organic matter in water reduces deeper
373
penetration of light (attenuation). The amount of dissolved organic matter (DOM) affects
374
exposure required sunlight mediated degradation of microorganisms. Suspended sediments,
375
presence of phytoplanktons, eutrophication exacerbates irradiance attenuation. A key
376
implication to these is that in humic-stained or eutrophic waters, sunlight mediated degradation
377
is less likely (Nelson et al. 2018).
378
Besides, bacteria have the ability to repair sunlight damage, and there is a chance for recovery
379
and regrowth if the injury by sunlight is sublethal (Nelson et al. 2018). Worse still, some
380
nosocomial pathogens (Serratia marcescens, Pseudomonas putida, and Stenotrophomonas) are
381
resistant to solar radiation (Glady-Croue et al. 2018). ARBs are more resistant to inactivation
382
during solar irradiation than non-resistant bacteria (Al-Jassim et al. 2017). Furthermore, a study
383
investigating the effect of sunlight on the decay of faecal indicator bacteria (FIB) in freshwater
384
demonstrated the minimal effect of sunlight on the survival of FIBs in freshwater (Korajkic et al.
385
2019). Bacterial photostress is complex and is determined by multiple environmental stressors.
386
For instance, higher concentrations of dissolved oxygen in the environment is required to
387
improve photoinduced damage of bacteria (Mcclary et al. 2017).
13
388
Sunlight-mediated pathogen inactivation does not guarantee a reduction in overall risk to
389
pathogen dissemination and the geographic kinds of human pathogens vary with distribution
390
and biological activity of UV radiation may change (Boehm et al. 2018).
391
Even if ARBs are effectively inactivated by sunlight, their DNA which contain ARG can still be
392
intact and capable of transferring resistance to non-resistant strains (Anyaduba 2016).
393
Information on the ability of sunlight to degrade ARGS is still scarce, but studies provide
394
evidence of monochromatic UVC radiation (UV254) preventing several ARGs from converting
395
competent non-resistant recipient bacteria into conforming resistant phenotypes (Nobuo and
396
Ikeda 1969; Setlow 1977; Chang et al. 2017). It is well established that UV can induce DNA
397
damage and leads to DNA base lesions such as pyrimidine (6-4) pyrimidone adducts [(6-4) and
398
cyclobutane-pyrimidine dimers (CPDs) photoproducts] (McKinney and Pruden 2012; Destiani et
399
al. 2018). A significant rate of UV–induced ARG degradation is documented (Destiani et al.
400
2018; Yoon et al. 2018). Also, exposure to artificial UV light eliminated plasmid-borne resistance
401
gene (Schuch and Menck 2010; Yoon et al. 2018). However, the plasmid-borne ARG that was
402
eliminated by UV is repairable during transformation by competent bacterial cells within the
403
environment (Nelson et al. 2018). The significant repair was observed in E. coli recipient strain
404
(DH5α) (Yoon et al. 2018). Also, ARGs degradation is slow (Nelson et al. 2018). Therefore,
405
sunlight mediated degradation does not completely eliminate ARGs in the freshwater
406
environment. The ARGs can be taken up by a competent bacteria, and incorporated into the
407
bacteria genome even when the original donor ARB cell is absent (Yoon et al. 2018). The rate at
408
which HGT occurs in the aquatic environment is not known quantitatively and is important to
409
determine whether HGT is a sink or source of ARGs in the aquatic environment.
410
Dilution
411
Dilution means the reduction of the concentration of a substance as it is dissolved in a larger
412
volume of a solvent (in this case water) (Lasagna et al. 2016). In aquatic environments dilution of
413
pollutants such as eDNA and microorganisms depends on currents, dynamic hydrological
414
processes and the amount of pollutants received (Floehr et al. 2013; Baldigo et al. 2017; Collins
415
et al. 2018). Where the body of water is large, with water current and strong tide the ARBs and
416
ARGs in water will quickly dilute and scatter (Foote et al. 2012). In addition, any substance
417
present in flowing water moves with the current. The dynamic hydrological processes include 14
418
advection and diffusion (Van Genuchten et al. 2013; Molz 2015). Advection refers to the
419
movement of objects suspended or dissolved in water along with the bulk flow, such as when a
420
river is flowing down a stream. Diffusion is the net motion of objects from a place of high
421
concentration to a place of low concentration, and therefore the substance spreads out in the
422
river. In terms of the amount of pollutants received, the assimilating capacity of the freshwater
423
can only cope with a certain part of the pollutants.
424
Dilution plays a substantial role in the reduction of the level of pollutants in water (Foote et al.
425
2012; Lasagna et al. 2013). It has been suggested as a process that is ever-present and not
426
influenced by either biological and chemical conditions (Lasagna et al. 2013). However, while
427
dilution reduces the concentration of a contaminant, it does not eliminate it from the system. In
428
agreement with the law of conservation of mass, the pollutants in freshwater become a part of
429
the hydrological cycle that is separated through adsorption and consequently become
430
dangerous wastes that will become of grave concern overtime(Sharma and Ahmad 2014).
431
Persistence of antibiotic resistance in the bacterial population
432
The presence of high concentrations of antibiotics in the environment causes direct selection for
433
resistance markers (Andersson and Hughes 2011). Therefore, freshwater bodies receiving
434
pharmaceutical waste streams or runoffs from agricultural farmlands can become polluted with
435
high concentrations of antibiotics, enabling selective pressure that encourages the preservation
436
and permanency of antibiotic resistance. Antibiotics, even at low concentrations can select for
437
resistant bacteria in the environment (Gullberg et al. 2011; Alm et al. 2014). Resistance becomes
438
permanent in bacterial populations under certain environmental conditions, even where
439
antibiotic selective pressure is low or absent for a significant period. Resistance can persist by (i)
440
compensatory mutations that reinstate fitness without loss of resistance; (ii) the incidence of
441
infrequent cost-free resistance mutation (iii) genetic association and co-selection between
442
resistance mutations and extra selected genetic markers (e.g. resistances or virulence factors)
443
(Andersson and Hughes 2011). Compensatory evolution lessens costs and allows maintenance
444
of resistance even without selective pressure.
445
Nevertheless, antibiotic resistance is not always the primary purpose of antibiotic resistance
446
genes. It has been suggested that some genes code for functions that confer a selective benefit 15
447
in the natural environment (Alm et al. 2014). This explained why Shwanella oneidensis MR-I
448
isolated from pristine sediments devoid of pharmaceutical impact possessed a Mex system
449
capable of antibiotic efflux, improving the fitness of the bacteria in the sediments (Groh et al.
450
2007). If the freshwater environment contains other substances, including anti-microbial
451
peptides released by other microbes, metals and organics, then Shwanella oneidensis can
452
benefit for common efflux pumps that are also applied for resistance to antibiotics (Alm et al.
453
2014). In terms of compensatory mutation, the fitness cost that is accompanying antibiotic
454
resistance is reduced over time as the bacteria become accustomed to the environment, and so
455
the pressure to lose the resistance phenotype is reduced (Moore et al. 2000). In terms of co-
456
selection, it is most likely that the resistance plasmids bear additional genes coding for proteins
457
that improve colonization, enabling alternative carbon utilization or improved nutrient uptake.
458
ARGs that have become fixed in a bacterium are difficult to eliminate (Andersson and Hughes
459
2011).
460
Therefore, having elucidated the processes that are involved in propagation and persistence of
461
antibiotic resistance in bacterial populations, as well as several mechanisms that contribute to
462
the stability of antibiotic resistance within the bacterial populations, it can be deduced that that
463
the release of antibiotics into freshwater environments contribute to the selection for resistance
464
within the bacterial population. Also important, one can deduce that antibiotic resistance
465
mechanisms will persist in freshwater systems in the presence of other substances that are not
466
antibiotics because the bacteria use mechanism that is similar for antibiotic resistance to survive
467
in the presence of substances that are toxic to them.
468 469
Investigating the potential transfer of antibiotic resistance in freshwater bodies
470
The prevalence of ARB and ARGs in freshwater environments implies that they are hotspots for
471
possible ARG dissemination. ARGs can be transferred through HGT mechanisms, including
472
transduction, conjugation, and transformation. In hotspots environments, different genetic
473
structures bearing ARGs are usually detected in addition to the ARGs themselves (Lupo et al.
474
2012). There are various types of mobile genetic elements, such as integrons, transposons,
475
plasmids, bacteriophages, as well as a combination of them. In particular, integrons are useful
476
targets for the detection of possible ARG transfer and spread because (i) they are one of the 16
477
simplest elements that participate in the transfer (mobility) of gene cassettes (ii) all integrons
478
have a common structure (iii) integrons can be linked to other mobile genetic elements, and (iv)
479
they can efficiently trap ARGs. Integrons can acquire, express and exchange ARGs that are
480
embedded in gene cassettes (GCs). An integron can be defined based on its structural
481
components. An integron is composed of 3 elements, namely intl gene, which encodes an
482
integrase, site attl specific for recombination and a promoter attI (Heuer et al. 2004; Cambray,
483
et al. 2010; Rizzo et al. 2013). Generally, GCs have an open reading frame that is coupled to an
484
attC site that is integrated or cut from the functioning platform by a site-specific recombination
485
mechanism that catalyze the intl integrase (Rizzo et al. 2013).
486
The two major groups of integrons are resistance integrons (RIs) and chromosomal integrons
487
(Cis). Cis is found on the chromosome of many bacterial species (Cambray, et al. 2010). RIs have
488
been investigated in several gram-negative bacteria, and few gram-positive bacteria (Nandi et
489
al. 2004; Xu et al. 2010). RIs are carried on mobile genetic elements, including plasmids and
490
transposons, which facilitate their transfer among bacteria. RIs are classified into five groups
491
based on the sequence of amino acid of the intl protein (Cambray, et al. 2010). Classes 1, 2 and
492
3 are the most studied. Classes 1 and 2 are frequently observed in environmental samples,
493
human and bacterial isolates (Stokes and Gillings 2011). Class 1 RIs are extensively described
494
and mostly encountered in multidrug-resistant bacteria. Only four class 3 RIs have been studied
495
in environmental Delftia isolates and clinical Enterobacteriaceaen strains (Xu et al. 2007).
496
Over 130 GCs encode resistance to nearly all antibiotics families such as aminoglycosides,
497
chloramphenicol, fosfomycin, lincosamides, β-lactams, rifampicin, macrolides, and quinolones. It
498
is because GC bore on integrons encodes resistance for a wide range of antibiotics that targeting
499
integron structures will provide an overall perspective on antibiotic resistance and the spread of
500
ARGs in the environment (Kristiansson et al. 2011). Fortunately, a qPCR method that targets
501
three main classes of integrons in DNA from complex matrices has been described (Barraud and
502
Ploy 2011). Thus, RIs can be applied in tracking the occurrence of resistance genes within the
503
environment. Furthermore, the presence of RIs can be used to indicate the acquisition of
504
antibiotic resistance genes. The prevalence of the Class 1 integron-integrase gene (intl1) has
505
been recommended as a reliable proxy for the occurrence of pollution by anthropogenic sources
506
(Gillings et al. 2015). Studies that show the occurrence of integron-bearing drug-resistant 17
507
bacteria in freshwater environments exists. In a study by Koczura et al. (2015), integron‐carrying
508
multidrug resistant coliform bacteria bearing virulence genes was observed in recreational lakes.
509
The presence of E. coli strains bearing Class 1 and 2 integrons isolated from freshwater in
510
Australia has been reported (Sidhu et al. 2017). A high prevalence of E.coli bearing class
511
1integrons was observed in the Minjian River in China (Chen et al. 2011).
512
Besides the class 1 integrase gene (intI1, transposase gene (tnpA) is another marker for mobile
513
genetic elements, that are important in the dissemination of resistance (Zhu et al. 2013; Gillings
514
et al. 2015; Hu et al. 2016). tnpA genes were detected in groundwater samples (Szekeres et al.
515
2018). However, intI1 responds quicker to environmental stressors than tnpA and so are
516
proposed as a better proxy for anthropogenic pollution and resistant (Gillings et al. 2015;
517
Szekeres et al. 2018)
518
In practice, however, it is very difficult to deduce the contribution of each and every
519
phenomenon to ARG transfer in the environments. In most cases, a high prevalence of ARB or
520
ARG does not show that gene transfer has happened. In most cases, resistance emergence may
521
be suggested to be due to horizontal gene transfer using ex post facto evidence (Rizzo et al.
522
2013). The major evidence to support ARG transfer include (i) the occurrence of a link between
523
the ARG acquired and mobile genetic elements, including phages, plasmids, transposons
524
(Partridge 2011), (ii) an observation of missing synteny between the DNA that is acquired and
525
the insertion site on the host (Dobrindt et al. 2004; Miriagou et al. 2006; Deurenberg and
526
Stobberingh 2008), (iii) a lack of resemblance between the phylogeny of the resultant host and
527
the supposedly transferred gene (Sørensen et al. 2005; Lal et al. 2008).
528
530
Evidence for the transfer of antibiotic resistance in freshwater environments
531
Direct evidence of transfer is usually scarce, and this is possible because it is difficult to study
532
such a transfer from a donor bacterium to potentially numerous resident bacteria that inhabit
533
complex environments (Sørensen et al. 2005; Rizzo et al. 2013). Generally, several antibiotic
534
resistance determinants that occur in clinical isolates are situated on mobile genetic elements
535
(MGE). This allows their horizontal transfer to other strains (commensals, pathogens, as well as
529
18
536
environmental) or even between bacteria from different taxa. Laboratory microcosm studies
537
are usually applied to demonstrate the transfer of plasmids to susceptible cells at population
538
densities and environmental conditions that are similar to what the bacteria encounter in the
539
natural freshwater environment (Alm et al. 2014). Conjugative plasmids isolated from multi-
540
resistant E.coli strains belonging to the IncP incapability group were transferred to Pseudomonas
541
fluorescens and Aeromonas sp. (Laroche-Ajzenberg et al. 2015). However, microcosm studies
542
show only transfer of a given mobile genetic element (mostly plasmids) and not all the genetic
543
exchanges that can potentially happen in complex microbial communities.
544
Nevertheless, examples of indirect evidence of transfer exists: (i)simultaneous detection of
545
similar resistance genes in pathogens, commensals and environment following the introduction
546
of a given antibiotic in clinics, (ii) increase in the number of bacterial antibiotic resistance genes
547
within the phage genome of the biomes of natural environments, anthropogenic environments
548
and in microbial communities of animal or humans (Muniesa et al. 2013), as well as (iii)
549
concurrent detection of mobile genetic elements positive pathogens in animals (i.e. migratory
550
birds) and their habitat. These are all examples of indirect evidence of transfer (Wu et al. 2018).
551
There are other indirect evidence of the transfer of antibiotic resistance in freshwater
552
environments. In freshwater bodies enteric bacteria are mostly found (Hu et al. 2008; Hoa et al.
553
2011; Suzuki et al. 2013), and they have very high potential to survive over time, long enough
554
for HGT to occur between the enteric bacterial community and aquatic environment (Vital et al.
555
2008). Besides, high species diversity that bacteria that occur in freshwater encourages HGT
556
(Suzuki et al. 2008). Furthermore, in an event of turbulence, faster ARG transfer through HGT
557
can be encouraged by mixing among the bacterial community (Andrup and Andersen 1999).
558 559
Transmission of antibiotic resistant bacteria and resistance genes from the freshwater environment to humans
560
Knowledge of transmission of antibiotic resistant bacteria and resistance genes from the
561
environment to humans is still scarce. A pathogen in a freshwater system may come in contact
562
with a strain that is resistant to antibiotics, for a period long enough for HGT to occur. For
563
freshwater bodies used for recreation, one can acquire an antibiotic resistant pathogen. Once
564
within human system, internal transfer within intestinal microbiota is also possible (Alm et al.
19
565
2014). Laurens et al. (2018) provide an insight into transmission from the environment to
566
humans. They investigated a human case of bacteraemia triggered by IMI-2 carbapenemase-
567
producing Enterobacter asburiae after exposure to river water. The authors applied Pulsed-field
568
gel electrophoresis (PFGE) to compare environmental and clinical bacterial strains. They also
569
used PFGE to determine the blaIMI-2 carbapenemase gene location. Thereafter, they applied
570
fingerprinting technique, 16S rRNA gene PCR–temporal temperature gel electrophoresis to
571
compare the patients' microbiota with those of the bacterial community of the river water to
572
which the patients were exposed. Their result indicated the same plasmidic blaIMI-2 gene was
573
carried in both E.asburiae causing the bacteraemia and that detected in river water a month
574
later. Both river and clinical strains displayed similar PFGE profiles. The patient’s microbiome of
575
carbapenem-resistant bacteria persisted and was autochthonous within the river community (E.
576
asburiae, Pseudomonas fluorescence and Aeromonas veronii). Laurens et al. (2018) then
577
hypothesized that if antibiotic resistance producing strains persisting in various geographic
578
locations are similar to the clinical isolates, then the antibiotic resistant strain may represent a
579
part of the natural reservoir of the resistance, and could be a vehicle for transferring resistance
580
between aquatic environment and human or animal. In addition, they argued that the flanking
581
of the blaIMI-2 gene initially described in E. asburiae by transposable elements on a conjugative
582
plasmid has the potential for the dissemination of gene among bacteria mainly the
583
Enterobacteriaceae (Rotova et al. 2017; Laurens et al. 2018).
584
With regard to ARBs, they can contaminate drinking water treatment plants and distribution
585
systems through freshwater. Drinking water treatment facilities source their raw water from
586
surface water including water from rivers and dams. The activities in the catchment area, as
587
well as the concomitant run-off, determine the level of pollution in surface water. Antibiotic
588
resistant pathogens can enter surface water from human waste (e.g. septic tank and sewage),
589
animal waste (e.g. animal dropping) and intensive farming practices (e.g. dairying and feedlots).
590
Insufficient treatment of water, including a failure to manage turbidity, and inadequate
591
chlorination can lead to pathogens being distributed through municipal water supply (Schwartz
592
et al. 2003; Nadiabartholomew et al. 2014). Schwartz et al. (2003) investigated wastewater,
593
surface water, and drinking water within one municipal system for the presence of resistant
594
bacteria and resistance genes. The authors detected vanA resistance gene without the
595
corresponding detection of bacteria (enterococci) in biofilms sampled from WWTP, wastewater 20
596
effluent, the effluent receiving river and the drinking water supply using the river as source
597
water. The result of the study by Schwartz et al. (2003) thus indicates a possible gene transfer to
598
indigenous drinking water bacteria.
599 600
Mitigation strategies for freshwater environment antibiotic resistance development
601
Antibiotic resistance spreads rapidly, such that even the smallest use of antibiotics significantly
602
increases the development and spread of antibiotic resistance. Limiting antibiotic use and
603
restricting ARG dissemination are the two main well-known methods to mitigate resistance
604
dissemination (Vikesland et al. 2017). Wastewater treatment plants (WWTPs) and hospital
605
effluents are globally accepted as major contributors of ARBs and ARGs in receiving aquatic
606
environments (Czekalski et al. 2012, 2014, 2015; Devarajan et al. 2016). More so, livestock
607
farming and other discharges from urban activities are pollutant sources of ARGs in freshwater
608
bodies, impacting on the resistance of microorganisms within the environment (Amy Pruden et
609
al. 2006; Zhang et al. 2013). Therefore, intervention could target these sources of ARGs in the
610
environment. Strategies to mitigate freshwater antibiotic resistance development could include
611
both traditional and inventive public health approaches, such as secondary treatment of
612
wastewater, providing and implementing standards and guidelines relating to discharges and
613
effluents quality and simple hygiene practices that eradicate or lessen the risk of contamination.
614
Creating methods that are more efficient for the treatment of wastewater from domestic,
615
hospital and industries that contain antimicrobial agents and ARGs before discharge into
616
freshwater is necessary. Disinfection processes that are used to eliminate pathogens during
617
wastewater treatment, including ozonation or chlorination are inadequate in destroying all the
618
genetic material in wastewater under most of the present conditions of operation (Xi et al.
619
2009; Dodd 2012; McKinney and Pruden 2012; Li et al. 2016; Pak et al. 2016; Chang et al. 2017).
620
More so, ozonation and chlorination can select for antibiotic resistance. In order to restrict ARGs
621
dissemination, additional extensive degradation of ARGs in wastewater is critical to address
622
antibiotic resistance. If traditional disinfection processes of wastewater treatment is
623
unavoidable, it is critical that bacteria from the system be eliminated before discharge (Pruden
624
et al. 2013).
21
625
Reducing antibiotics usage in livestock farming will decrease selection pressures and
626
consequently will reduce the preservation of ARGS within the host so as to attenuate resistant
627
strains over time (Bengtsson-Palme and Larsson 2016; Vikesland et al. 2017).
628
antibiotics usage in livestock farming can be achieved by sustaining good animal health and
629
reducing the incidence of disease through the improvement of the conditions under which the
630
animals are bred, and the use of substitutes to antibiotics (Thanner et al. 2016; Tullo et al.
631
2019). Additionally, distancing livestock farming activity from freshwater systems is another
632
preventive measure of antibiotic resistance spread.
633
At a policy level, in order to curtail the spread of antimicrobial resistance from these sources
634
standards on the concentration of antibiotics, which are selecting agents for ARBs, in
635
wastewater treatment plants and hospital effluents to be discharged into nearby aquatic
636
environments must be established. It may be valuable to enforce more restrictions on the use of
637
those antibiotics that persist in the environment, such as fluoroquinolones. Pharmaceutical
638
industries must practice good manufacturing practices that include consideration of the
639
environment, as this could be of benefit (Pruden et al. 2013). So far, it is understood that
640
endorsing regulations on management of effluents from WWTP and even pharmaceutical
641
industries is still a challenge because the cooperation, agreement and enforcement by a large
642
number of stakeholders will be required (Pruden et al. 2013).
643
Conclusion
644
Antibiotics resistant bacteria and resistance genes are prevalent in freshwater environments,
645
and anticipative natural processes of degradation and dilution are not able to completely
646
eradicate them. ARBs are resistant to inactivation through microbial and sunlight degradation,
647
and even if they are effectively inactivated by sunlight, their ARG-containing DNA can still be
648
intact and capable of transferring resistance to non-resistant strains. Rivers with high water
649
quantity can dilute ARGs, but the assimilating capacity of the freshwater can only cope with a
650
certain quantity of pollutants. The ARBs found in freshwater environments might constitute part
651
of the natural reservoir of antibiotic resistance and can act as a vehicle between freshwater
652
bodies and human microbiota. The evidence provided in this study sustains that direct
653
transmission of indigenous freshwater ARBs to humans as well as their transitory insertion in the
654
microbiota can occur. This study provides a baseline knowledge that is fundamental for any risk 22
Reducing
655
assessment study of freshwater bodies with regard to ARGs and ARB. It would be prime if
656
antibiotic usage is reduced to mitigate resistance dissemination.
657
hundreds of millions of people lack access to treated water that is safe for consumption and so
658
relies on freshwater resources for drinking, sustaining crops through irrigation and providing
659
food in the form of fish. The consumption of water that is polluted with ARBs and ARGs or food
660
irrigated with the contaminated water may facilitate the dissemination of antibiotic resistance
661
to humans.
662
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Table 1 Clinically relevant ARBs detected in freshwater bodies ARB group
Extended-spectrum βlactamase producing Enterobacteriaceae
ARG carried
Specie
No
No
of isolates
positive
E. coli
9
6
K. pneumoniae
1
1
E. cloacae
1
1
E. coli
38
38
K. pneumoniae
2
2
E. cloacae
1
1
E. coli
5
1
C. frendii
2
1
E. coli
5
2
K. pneumoniae
2
1
blaTEM−1
K. pneumoniae
2
1
blaCMY−2
E. coli
5
1
blaCTX-M-15
blaCTX-M-14-like
blaSHV
blaSHV−1 in association with blaCTX−M
1
Type freshwater
of
References
Stream
Miyagi and Hirai (2019)
River water
Ye et al. (2017)
C. frendii
2
1
E. coli
5
1
blaSHV
E. coli
74
3
blaCTX−M
E. coli
74
71
blaIMI‐2
Enterobacter asburiae
7
Almost all of the isolates in this study had at least one of the genes (blaTEM, blaCTX-M, blaSHV and blaOxA)
E. coli
167
Klebsiella pneumoniae
114
Citrobacter freundii,
9
Enterobacter cloacae
6
blaSHV−1, blaTEM−1, simultaneously
2
and
blaCTX−M−65
River
Zurfluh et al. (2013)
7
River
Harmon et al. (2018)
Not applicable here
River
Chen et (2010)
al.
Citrobacter koseri
4
Salmonella ssp. Arizonae
3
choleraesuis
Serratia liquefaciens
3
Pantoea spp.
1
blaCTX-M-14
E. coli
30
10
River
Dhanji (2011)
et
al.
blaVEB
Aeromonas spp
29
11
River
Girlich (2011)
et
al.
blaSHV-12
29
10
blaPER-1
29
3
blaTLA-2
29
1
blaGES-7
29
1
Not applicable
Not applicable
River
Tacão et (2012)
al.
blaCTX-M, blaTEM, blaOXA-21
A. hydrophila
integrase genes intI1
Escherichia coli Pseudomonas sp Acinetobacter sp
3
Carbapenemase-producing Enterobacteriaceae
methicillin resistant Staphylococcus aureus
blaKPC-2
E. coli
3
3
K. pneumoniae
1
1
Enterobacter cloacae
1
1
blaIMI-2
Enterobacter cloacae
1
1
blaKPC-2 and blaVIM-1
Klebsiella oxytoca
2
2
mecA gene
Staphylococcus aureus
22
mecA
Staphylococcus aureus
mecA
Staphylococcus aureus
River sediments
Piedra-Carrasco et al. (2017)
22
freshwater beach
Levin-Edens et al. (2012)
698
12
River
Hatcher et al. (2016)
70
24
Freshwater recreational beach
Thapaliya et al. (2017)
70
15
32
27
River
Novais (2005)
vanB
32
4
vanC1
32
1
PVL
Vancomycin Enterococci
resistant
vanA
Enterococci faecalis
4
et
al.
vanB
E. faecalis
Not applicable
Not applicable
vanC
333
166
Van C1
333
10
vanC2/C3
E. casseliflavus
333
164
VanC
Enterococcus spp
216
61
216
1
Van C1
Lata et (2009)
al.
River
Nishiyama et al. (2017)
River
Nam et (2013)
al.
al.
vanA
Enterococcus faecium
20
1
River
Morris (2012)
vanC1
Enterococcus casseliflavus/gallinarum
Not applicable
Not applicable
River
Roberts et al. (2009)
vanC2/3
Enterococcus casseliflavus/gallinarum
5
et
49.7
River
40.3
50.9
39.5
32.8
16S
rRNA Devarajan
gene
sediments
normalized log copy
6
et al. (2016)
Table 1Reported prevalence of antibiotic resistant bacteria in freshwater ecosystems Type of freshwater
Antibiotics
Pond
Carbapenem
% Antibiotic resistant bacteria
References
0.04 Harmon et al. (2019)
0.10 Pond
0.01 Pond
0.05 Dam
0.00 Spring
0.00 Lake
3.24 Lake
0.01 Lake
0.09 Lake
0.02 Natural Pool
0.45 River
Tacão et al. (2015)
77.27 Lake
Penicillin
Pang et al. (2015)
64.55 Ampicillin
34.55 Cephalothin
14.55 Chloramphenicol
77.27 Tetracycline Rifampicin
1.00
1
Type freshwater
of
Antibiotics
% Antibiotic resistant bacteria
References
6.6–21.0 River
Ampicillin
Ash et al. (2002) and Aubron et al. (2005)
10.4–25.7 22.6–24.1 34.5–38.4 26.9 59.2 26.4 13.7–34 10.1–36.6 4.9–52.5 5.9 19.7–23.7 6.7–73.0 6.1–21.5 20.0–53.0 12.4–20.0
2
Type of freshwater
Antibiotics
%
References
Antibiotic resistant bacteria
15.7 River
Ampicillin
Ash et al. (2002) and Aubron et al. (2005) 32 3.5–33.9 3.9 17.0–25.0
22.5
93 Cephalothin 96 94 91 8 86 25 73 67 77
3
Type freshwater
of
Antibiotics
% Antibiotic resistant bacteria
References
78 River
Cephalothin
Ash et al. (2002) and Aubron et al. (2005)
36 96 86 38 98 94 96 5 Cefotaxime
6 13 9 25 12 0 5
4
Type of freshwater
Antibiotics
%
References
Antibiotic resistant bacteria
9 River
Cefotaxime
Ash et al. (2002) and Aubron et al. (2005) 20 21 4 0 8 0 16 2 0 4 7 39 2
Ceftazidime 0
5
Type of freshwater
Antibiotics
River
Ceftazidime
% Antibiotic resistant bacteria
References
1 Ash et al. (2002) and Aubron et al. (2005) 6 0 0 8 5 2 0 9 0 2 0 0 0 0 0
6
Type of freshwater
Antibiotics
%
References
Antibiotic resistant bacteria
5 River
imipenem
Ash et al. (2002) and Aubron et al. (2005) 0 0 10 0 0 4 1 0 0 6
68 Amoxicillin+clavalanic 8 acid 10 37 36
7
Type of freshwater
Antibiotics
% Antibiotic resistant bacteria
River
0 Amoxicillin+clavalanic 24 acid
References
Ash et al. (2002) and Aubron et al. (2005)
8
Table 1 Abundance of antibiotic resistance genes detected in freshwater bodies (sul, tet, aa, bla, flo as well as erm and mef signify sulfonamide, tetracycline, aminoglycoside, β-lactam, chloramphenicol as well as Macrolide-Lincosamide-Streptogramin) Type of fresh-water body River sediments
Type of ARG detected sul1
Quantity detected
Unit of measurement
10− 6 - 10− 3
Gene copy numbers normalized to bacterial 16S rRNA gene copy number
References
Amy Pruden et al. (2006); Pei
et
al.
(2006);
Storteboom et al. (2010)
Lake sediments
blaSHV blaTEM blaCTX-M blaSHV blaNDM aadA blaSHV blaTEM blaCTX-M blaSHV blaNDM aadA sul1
0–6.4 4.7–6.1 1.7–6.1 0–6.4 3.5–5.1 2.2–5.2 39.5 49.7 50.9 39.5 32.8 40.3 7.3 × 10− 1 ± 9 × 10− 2
sul2
4 × 10− 2 ± 3 × 10
sul1 tet(B) blaSHV
2.2 × 109 1.5 × 106 1.4 × 10–6 to 8.9 × 10–4
blaTEM blaCTX-M blaSHV blaNDM aadA
2.6 × 10-5 to 2.8 × 10–3 3.2 × 10-6 to 1.7 × 10–3 1.4 × 10–6 to 8.9 × 10–4 6.07 × 10–6 and 1.2 × 10–6 from 1.5 × 10–5 to 1.1 × 10–3
16S rRNA gene normalized log copy
Devarajan et al. (2016)
16S rRNA gene normalized log copy
Devarajan et al. (2016)
Gene copy number normalized to bacterial 16S rRNA gene copy numbers
Czekalski et al. (2012)
ARG copies g−1 wwt
Czekalski et al. (2014)
16S rRNA gene normalized log copy
Devarajan et al. (2015)
Table 2 continued. Abundance of antibiotic resistance genes detected in freshwater bodies (sul, tet, aa, bla, flo as well as erm and mef signify sulfonamide, tetracycline, aminoglycoside, β-lactam, chloramphenicol as well as Macrolide-Lincosamide-Streptogramin)
Type of fresh-water body Lake water
Sediments and river water
Type of ARG detected sul1
Quantity detected
Unit of measurement
References
1.5 × 10−3 - 2.1 × 10
Gene copy numbers normalized to bacterial 16S rRNA gene copy numbers
Czekalski et al. (2015)
sul2 blaTEM
Up to 3.4 × 10−3 Both samples analysed were positive for these ARGs.
log10 GC/mL or g) of each ARG in the bacterial (BAC)
Piedra-Carrasco et al. (2017)
qnrA qnrS mecA blaCTX-M blaSHV blaNDM Groundwater sul1 sul2 tet(A) tet(C) tet(O) tet(W) blaSHV floR ermA mefA
The abundances of copies/16S rRNA gene these ranged from copies 6.61 × 10−7 to 2.30 × 10−1
Szekeres et al. (2018)
Figure 1 A depiction of the potential origin, spread, transmission and degradation/dilution of ARBs and ARGs
Freshwater environment is a reservoir where antibiotics impact microorganisms Carbapenem resistant bacteria are less prevalent in freshwater environments Different ARGs, sul1, sul2, tet(B), aadA, blaTEM, blaCTX-M, blaSHV, blaNDM are detectable in Freshwater. Natural processes of degradation and dilution not able to eradicate ARBs and ARGs in freshwater. The ARBs can act as a vehicle between freshwater bodies and human microbiota. The major contributors of ARBs and ARGs are points for mitigating antibiotic resistance development.
This study provides knowledge that is fundamental for any risk assessment study estimating the actual risk of antibiotic resistance dissemination from freshwater reservoirs.
The authors declare that there is no conflict of interest.