Wastewater Treatment Effluent Reduces the Abundance and Diversity of Benthic Bacterial Communities in Urban and Suburban Rivers

Bradley Drury; Emma Rosi-Marshall; John J Kelly

doi:10.1128/AEM.03527-12

. 2013 Mar;79(6):1897–1905. doi: 10.1128/AEM.03527-12

Wastewater Treatment Effluent Reduces the Abundance and Diversity of Benthic Bacterial Communities in Urban and Suburban Rivers

Bradley Drury ^a, Emma Rosi-Marshall ^b, John J Kelly ^a,^✉

PMCID: PMC3592216 PMID: 23315724

Abstract

In highly urbanized areas, wastewater treatment plant (WWTP) effluent can represent a significant component of freshwater ecosystems. As it is impossible for the composition of WWTP effluent to match the composition of the receiving system, the potential exists for effluent to significantly impact the chemical and biological characteristics of the receiving ecosystem. We assessed the impacts of WWTP effluent on the size, activity, and composition of benthic microbial communities by comparing two distinct field sites in the Chicago metropolitan region: a highly urbanized river receiving effluent from a large WWTP and a suburban river receiving effluent from a much smaller WWTP. At sites upstream of effluent input, the urban and suburban rivers differed significantly in chemical characteristics and in the composition of their sediment bacterial communities. Although effluent resulted in significant increases in inorganic nutrients in both rivers, surprisingly, it also resulted in significant decreases in the population size and diversity of sediment bacterial communities. Tag pyrosequencing of bacterial 16S rRNA genes revealed significant effects of effluent on sediment bacterial community composition in both rivers, including decreases in abundances of Deltaproteobacteria, Desulfococcus, Dechloromonas, and Chloroflexi sequences and increases in abundances of Nitrospirae and Sphingobacteriales sequences. The overall effect of the WWTP inputs was that the two rivers, which were distinct in chemical and biological properties upstream of the WWTPs, were almost indistinguishable downstream. These results suggest that WWTP effluent has the potential to reduce the natural variability that exists among river ecosystems and indicate that WWTP effluent may contribute to biotic homogenization.

INTRODUCTION

Centralized wastewater treatment plants (WWTPs) are one of the most common systems for the treatment of domestic wastewater in the United States (1). In highly urbanized areas with high population densities, WWTPs can be large and numerous. For example, Cook County, IL, which includes the city of Chicago and is the second most populous county in the United States (www.census.gov), is serviced by seven WWTPs. One of these, the Stickney Water Reclamation Plant, is the largest activated sludge WWTP in the world, with a design capacity of 1.2 billion gallons per day (www.mwrd.org). WWTPs frequently discharge effluent water into lotic ecosystems, and in many cases, WWTP effluent makes up a significant proportion of the flow of the receiving water body (2). For example, treated municipal wastewater effluent is more than 70% of the annual flow in the Chicago Area Waterway System, which includes all segments of the Chicago River as well as the North Shore Channel (NSC) (3). Therefore, in highly urbanized areas like Cook County, IL, WWTP effluent represents a significant component of the water in lotic ecosystems.

Although WWTPs can be effective at reduction of biochemical oxygen demand (BOD) and pathogen load, it is impossible for the characteristics of the effluent to match the characteristics of the water in the receiving system. Therefore, the potential exists for WWTP effluent to significantly alter the physical and chemical properties of the receiving ecosystem. Numerous studies have documented the potential ecosystem effects of WWTP effluent, including increased nutrient loading (4) and eutrophication (5). Several previous studies have also examined the effects of WWTP effluent on bacterial populations within the water column (6–8), and some have demonstrated the ability of microorganisms contained in the effluent to persist in the water column of the receiving system (8). However, few studies have examined the potential effects of WWTP effluent on benthic microbial communities (9) despite the fact that bacterial numbers are generally much higher in freshwater sediment than in the overlying water (10) and despite the fact that benthic microbial communities are critical components of lotic ecosystems, as they contribute to organic matter decomposition, nutrient cycling, and bioremediation of a variety of pollutants. Several recent studies have presented evidence that WWTP effluent may impact the function and structure of sediment microbial communities. For example, Lofton et al. (11) previously reported a significant increase in denitrification rates in sediment samples collected downstream of a WWTP in Greensboro, NC. In addition, Wakelin et al. (9) previously used denaturing gradient gel electrophoresis (DGGE) to demonstrate that effluent from a small WWTP altered the composition of sediment bacterial communities in a small rural stream in Australia. However, we are not aware of any study that has examined the effects of WWTP effluent on sediment microbial community function and structure in a highly urbanized habitat within a major city.

We assessed the impacts of WWTP effluent on the size, activity, and composition of benthic microbial communities in lotic ecosystems in two distinct field sites in the Chicago metropolitan region: (i) a river in a highly urbanized area receiving effluent from a large WWTP and (ii) a river in a less densely populated suburban area receiving effluent from a much smaller WWTP. Our data demonstrate that although these WWTPs differed dramatically in size, they had remarkably similar effects on the chemical and biological properties of the receiving rivers, including increases in concentrations of inorganic nutrients, decreases in the population size of sediment bacterial communities, and shifts in the composition of sediment bacterial communities. The net effect of WWTP inputs was that two rivers that were distinct in chemical and biological properties upstream of the WWTPs were almost indistinguishable downstream. These results suggest that WWTP effluent may have the potential to reduce the natural chemical and biological variability that exists among river ecosystems.

MATERIALS AND METHODS

Field sites.

The North Shore Channel (NSC) was selected to represent a highly urbanized river. The NSC is a 7.7-mile-long canal that begins in the town of Wilmette, IL, and extends into the northeast section of the city of Chicago, IL. The canal was built in 1910 to bring water from Lake Michigan to the North Branch of the Chicago River. The NSC has an average discharge of 0.93 m³ s⁻¹ (http://waterdata.usgs.gov/) and a drainage area of 6,474 ha that is 63% residential, 16.7% commercial/industrial, 10% forest/open land, 5.4% institutional, and 3.5% transportation/utility (12). The NSC receives treated effluent from the North Side Water Reclamation Plant (NSWRP), an activated sludge plant that receives domestic wastewater from over 1.3 million people residing in a 141-square-mile area that includes part of the city of Chicago and the northern Cook County suburbs. The NSWRP has an average flow of 245 million gallons per day (MGD) and a design capacity of 333 MGD. The NSWRP treats wastewater with a series of physical and biological processes, and effluent is not disinfected prior to release (www.mwrd.org). Two sampling sites on the NSC were chosen, one approximately 925 m upstream of the input of effluent from the NSWRP and one approximately 50 m downstream of the effluent input.

The West Branch of the DuPage River (WBDR), located in DuPage County, IL, was selected to represent a suburban river. The WBDR has an average discharge of 0.62 m³ s⁻¹ and a drainage area of 32,900 ha that is 32.8% residential, 17.4% agricultural, 16.9% vacant, 11.2% forest/open land, and less than 4% industrial (13). The WBDR receives treated effluent from the West Chicago WWTP (WCWWTP), which is located in West Chicago, IL. The WCWWTP is an activated sludge plant that receives domestic wastewater from the towns of West Chicago and Winfield, IL. It treats 5 MGD and does not disinfect the effluent prior to release (www.westchicago.org). Two sampling sites on the WBDR were chosen, one approximately 275 m upstream of the input of effluent from the WCWWTP and one approximately 50 m downstream of the effluent input.

Sample collection.

Five replicate sediment samples and five replicate water samples were collected at each of the four sampling sites between August and September 2010. Each sediment sample consisted of a composite of 10 individual sediment samples collected from randomly selected sections along the stream reach. Sediment samples were collected by using a Petite Ponar sampler (Wildlife Supply Company, Saginaw, MI), and large debris was removed by hand. Sediment samples were stored in sterile 400-ml canning jars (Ball Corporation, Muncie, IN). Water samples were collected into sterile, precleaned, 1-liter amber glass jars (Thermo Scientific, Waltham, MA). All sediment and water samples were stored on ice for transport back to the laboratory.

Sample characteristics.

Dissolved organic carbon (DOC) in water samples was measured on a Shimadzu 5050 TC analyzer as described previously (14). Ammonium, nitrate, and phosphate concentrations in water samples were determined with a Lachat QuikChem 8000 instrument by the phenate method (method 10-107-06-1-J; Lachat Instruments, Milwaukee, WI), the cadmium diazotization method (method 10-107-04-1-C; Lachat Instruments), and the phosphomolybdate method (method 10-115-01-1-M; Lachat Instruments), respectively. Sediment organic material was measured by loss on ignition at 500°C (15).

Microbial respiration.

Respiration was measured for each sediment sample by using a standard method (16). Briefly, 10 ml of sediment was placed into a black high-density polyethylene 50-ml centrifuge tube (Cole-Parmer, Vernon Hills, IL) filled to the top (no headspace) with well water. Water temperature and initial dissolved oxygen (DO) were measured by using a YSI ProODO meter (YSI Inc., Yellow Springs, OH). Centrifuge tubes were capped, eliminating all air bubbles, and incubated at room temperature (25°C) in the dark for 2 h, after which final DO was measured and respiration rates were calculated as mg O₂ consumed time⁻¹. Respiration rates were then normalized by sediment surface area and by total heterotrophic plate counts.

Heterotrophic plate counts.

Viable counts of heterotrophic bacteria were conducted for each sediment sample by using a standard plate count method (17). Briefly, 10 g of sediment was placed into a sterile 250-ml centrifuge bottle containing 90 ml of heat-sterilized potassium phosphate buffer solution. Samples were agitated for 30 min at 300 rpm by using a reciprocal shaker (New Brunswick Scientific, Edison, NJ). Samples were allowed to settle for 5 min, 1 ml of supernatant was serially diluted 10-fold to 10⁻⁵, and 100 μl of each dilution was plated onto soy extract agar (Becton, Dickinson and Company, Sparks, MD) plates containing 100 mg liter⁻¹ cycloheximide (MP Biomedicals, Solon, OH) to inhibit fungal growth. Numbers of CFU were normalized based on grams of dry sediment.

Epifluorescence counts.

Direct counts of bacterial cells were performed by using a modified standard method (18). Cells were fixed by diluting sediment 1:50 in sterile DNA-free fixative solution (10 mM NaPO₄, 120 mM NaCl, 10 mM sodium pyrophosphate, 4% formaldehyde) (19) in a sterile 50-ml centrifuge tube. Samples were placed into an ultrasonic ice water bath (model 8845-30; Cole-Parmer, Vernon Hills, IL) and sonicated for 15 min at 60 Hz. Following ultrasonic treatment, samples were diluted 1:1,000, 1:2,000, and 1:4,000 in 0.2-μm-filtered deionized water. Two milliliters of each diluted sample was filtered in duplicate onto 0.2-μm Anodisc membrane filters (Whatman, Maidstone, United Kingdom) and stained with 100 μl of SYBR gold (Invitrogen, Carlsbad, CA). Cells were counted at a ×400 magnification by using an Olympus BH-2 fluorescence microscope (Olympus, Center Valley, PA). Cell numbers were normalized based on grams of dry sediment.

Tag pyrosequencing.

DNA was isolated from each of the sediment samples by using the UltraClean Soil DNA kit (MoBio Laboratories, Carlsbad, CA). Successful DNA isolation was confirmed by agarose gel electrophoresis. For tag pyrosequencing of bacterial 16S rRNA genes, extracted DNA was sent to the Research and Testing Laboratory (Lubbock, TX). PCR amplification was performed by using primers 530F and 1100R (20). Primer 530F was chosen in order to obtain sequences for the V4 hypervariable region, which has been shown to provide species richness estimates comparable to those obtained with the nearly full-length 16S rRNA gene (21). Sequencing reactions were performed by utilizing a Roche 454 FLX instrument (Roche, Indianapolis, IN) with Titanium reagents. Sequences were processed by using MOTHUR v.1.20.1 (22). Briefly, any sequences containing ambiguities or homopolymers longer than 8 bases were removed. The remaining sequences were individually trimmed to retain only high-quality sequence reads, and sequences were aligned based on comparison to the SILVA-compatible bacterial alignment database available within MOTHUR. Aligned sequences were trimmed to a uniform length of 250 bp, and chimeric sequences were removed by using Chimeraslayer (23) run within MOTHUR. Sequences were grouped into phylotypes by comparison to the SILVA-compatible bacterial alignment database available within MOTHUR, and chloroplast sequences were removed. After these pretreatment steps were completed, the data set included a total of 173,842 sequences for an average of 8,692 sequences per sample. Sequences were clustered into operational taxonomic units (OTUs) based on 97% sequence identity by using the average neighbor algorithm. The community compositions of the individual sampling sites were compared by using MOTHUR to calculate distances between sites based on the theta index (24) and to visualize the resulting distance matrix by using nonmetric multidimensional scaling (NMDS). The significance of differences in theta index scores between sites was assessed by AMOVA run within MOTHUR. MOTHUR was also used to calculate the Shannon diversity index (25) and the Chao1 richness estimator (26). SIMPER analysis run in Primer V.5 (Primer-E Ltd., Plymouth, United Kingdom) was used to identify the OTUs making the largest contributions to the variations between communities from each of the field sites.

Statistics.

All data were analyzed by two-way analysis of variance (ANOVA) based on land use (urban versus suburban) and location (upstream of effluent input versus downstream). Analyses were run by using Systat version 12 (Systat Software Inc., San Jose, CA), and P values of less than 0.05 were considered to be significant.

RESULTS

Effects of land use.

We examined the influence of land use on the two rivers by comparing data from the reaches above the wastewater effluent inputs. There was a significant effect of land use (urban versus suburban) on water column nutrient concentrations, with higher concentrations of DOC, nitrate, and phosphate and a lower concentration of ammonium in the suburban river (Table 1). There was also a significant effect of land use on the size of the sediment bacterial communities, as indicated by heterotrophic plate counts, with the suburban river having higher counts than the urban river (Fig. 1A). Direct epifluorescence counts of bacterial cells did not show this same trend, as there was no significant effect of land use (Fig. 1B). There was also no significant effect of land use on microbial community respiration (Fig. 2A). NMDS analysis of tag pyrosequencing data revealed a significant difference in community composition between the suburban and urban rivers at the upstream sites (P < 0.001) (Fig. 3). Sediment communities from the upstream sites of both rivers were dominated by proteobacteria, with proteobacterial sequences accounting for more than 50% of the sequences (Fig. 4). There was no significant difference in the overall abundance of proteobacterial sequences between the two sites (P = 0.180), but the urban upstream site showed a significantly higher relative abundance of Bacteroidetes sequences (P = 0.001) and a significantly lower relative abundance of Chloroflexi sequences (P < 0.001) (Fig. 4). SIMPER analysis of the pyrosequencing data indicated 16 OTUs that accounted for more than 20% of the variation in community composition between the suburban and urban upstream sites, and there were significant differences in the relative abundances of each of these OTUs between the two sites (Table 2).

Table 1.

Sampling site characteristics

Parameter	Mean value for site (SE)^a				P value by ANOVA
Parameter	Suburban upstream	Suburban downstream	Urban upstream	Urban downstream	Land use effect	Effluent effect	Interaction effect
Water column DOC (mg liter⁻¹)	6.652 (0.052)	5.782 (0.306)*	2.408 (0.085)	3.947 (0.072)*	<0.001	0.06	<0.001
Water column NH₄ (mg liter⁻¹)^b	0.060 (0.003)	BD*	0.138 (0.007)	0.236 (0.005)*	<0.001	<0.001	<0.001
Water column NO₃⁻ (mg liter⁻¹)^b	2.742 (0.140)	4.662 (0.492)*	0.232 (0.002)	4.696 (0.206)*	<0.001	<0.001	<0.001
Water column PO₄³⁻ (mg liter⁻¹)^c	0.268 (0.006)	0.466 (0.035)*	0.003 (0.000)	0.410 (0.019)*	<0.001	<0.001	<0.001
Sediment organic material (%)	8.70 (1.20)	1.58 (0.12)*	5.89 (0.43)	2.00 (0.21)*	0.216	<0.001	0.025

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Each data point is the mean (n = 5), with standard error values in parentheses. Asterisks indicate a significant effect of effluent input (P value) based on two-way ANOVA. BD, below the detection limit.

The limit of detection for NH₄ and NO₃⁻ was 0.02 mg/liter.

The limit of detection for PO₄³⁻ was 0.002 mg/liter.

Fig 1 — Heterotrophic plate counts (A) and direct bacterial cell counts (B) for sediments collected from rivers within two land use types (urban and suburban) from locations upstream and downstream of WWTP effluent inputs. Each data point is the mean (n = 5) ± standard error. Two-way ANOVA for plate counts demonstrated significant effects of land use (P = 0.005) and effluent input (P = 0.000) but no significant interaction effect (P = 0.607). Two-way ANOVA for direct counts demonstrated no significant effect of land use (P = 0.091) but a significant effect of effluent input (P = 0.001) and no significant interaction effect (P = 0.243).

Fig 2 — Community respiration normalized by surface area (A) and by bacterial cell numbers based on heterotrophic plate counts (B) for sediments collected from rivers within two land use types (urban and suburban) from locations upstream and downstream of WWTP effluent inputs. Each data point is the mean (n = 5) ± standard error. Two-way ANOVA for respiration normalized by surface area demonstrated no effect of land use (P = 0.572) or effluent input (P = 0.189) and no interaction effect (P = 0.176). Two-way ANOVA for respiration normalized by cell numbers demonstrated a significant effect of land use (P = 0.000) and effluent input (P = 0.000) and a significant interaction effect (P = 0.000).

Fig 3 — NMDS ordination of 16S tag pyrosequencing data comparing community structures of sediment bacterial communities collected from rivers within two land use types (urban and suburban) from locations upstream and downstream of WWTP effluent inputs.

Fig 4 — Phylotype analysis of 16S tag pyrosequencing data for sediment bacterial communities collected from rivers in two land use types (urban and suburban) from locations upstream and downstream of WWTP effluent inputs. The y axis represents the percentage of all sequences within a sample that were within the phylotype listed on the x axis. Each data point is the mean (n = 5) ± standard error. An asterisk indicates a significant effect of effluent input (P < 0.05).

Table 2.

Bacterial operational taxonomic units making the most significant contribution to variation between communities from suburban upstream and urban upstream sites

Operational taxonomic unit^c	Relative abundance (%)^a		P value^b	Contribution to variation (%)	Cumulative contribution to variation (%)	Taxonomic identification^d
Operational taxonomic unit^c	Suburban upstream	Urban upstream	P value^b	Contribution to variation (%)	Cumulative contribution to variation (%)	Taxonomic identification^d
Otu1	0.18	5.17	<0.001	3.55	3.55	Dechloromonas
Otu3	5.90	2.10	0.031	2.98	6.53	Crenothrix
Otu5	0.33	4.39	<0.001	2.89	9.41	Thiobacillus
Otu12	2.84	0.02	<0.001	2.01	11.42	Comamonadaceae
Otu16	0.93	2.41	0.001	1.05	12.47	Proteobacteria, unclassified
Otu24	0.37	1.72	<0.001	0.96	13.43	Alteromonadaceae
Otu11	0.10	1.41	0.002	0.94	14.36	Giesbergeria
Otu22	0.07	1.35	<0.001	0.91	15.27	Nitrospira
Otu37	0.13	1.29	<0.001	0.83	16.10	Caldilineaceae
Otu29	0.21	1.11	<0.001	0.64	16.74	Sinobacteraceae
Otu17	0.00	0.90	0.001	0.64	17.38	Sphingobacteria
Otu30	1.01	0.11	<0.001	0.63	18.01	Methylococcus
Otu35	1.04	0.22	0.004	0.58	18.59	Thiovirga
Otu43	1.06	0.25	0.013	0.58	19.17	Perlucidibaca
Otu6	0.07	0.84	<0.001	0.55	19.72	Deltaproteobacteria
Otu176	0.78	0.02	<0.001	0.54	20.26	Pseudomonas

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Each data point is the mean value (n = 5).

P value based on ANOVA comparison of suburban upstream and urban upstream samples.

OTUs were identified within a 16S tag pyrosequencing data set based on 97% sequence identity.

Taxonomic assignments were based on comparison to the SILVA-compatible bacterial alignment database.

Effects of effluent.

WWTP effluent had significant effects on water column ammonium, nitrate, and phosphate concentrations as well as sediment organic content in both rivers (Table 1). Specifically, WWTP effluent resulted in significant increases in water column nitrate and phosphate concentrations at both the urban and suburban sites and a significant increase in the water column ammonium concentration at the urban site. WWTP effluent also resulted in significant decreases in sediment organic material content at both the urban and suburban sites. WWTP effluent had significant effects on the population size of the sediment bacterial communities, as indicated by both heterotrophic plate counts (Fig. 1A) and direct counts (Fig. 1B). Specifically, WWTP effluent resulted in significant decreases in both plate counts and direct counts at both sites. In contrast, WWTP effluent had no effect on community respiration normalized by sediment surface area (Fig. 2A). However, the decreased bacterial population size (Fig. 1A and B) combined with similar respiration rates (Fig. 2A) suggests smaller populations respiring more on a per-cell basis. When we normalized respiration rates by cell counts (Fig. 2B), both sites downstream of WWTP effluent had higher per-cell respiration rates than the upstream sites. This effect was much more pronounced at the urban site.

NMDS analysis of tag pyrosequencing data indicated that WWTP effluent significantly changed the composition of the sediment bacterial communities at both the urban (P < 0.001) and suburban (P < 0.005) sites (Fig. 3). These analyses also indicated that WWTP effluent significantly reduced bacterial community diversity (Fig. 5A) and richness (Fig. 5B). When we compared the bacterial communities of the two sites below the effluent inputs (i.e., urban downstream versus suburban downstream), we found no significant differences in community composition (P = 0.982) (Fig. 3), diversity (Fig. 5A), or richness (Fig. 5B). In terms of broad bacterial phyla, tag pyrosequencing revealed that WWTP effluent significantly reduced the relative abundance of proteobacterial sequences within the sediment bacterial communities (Fig. 4). This overall decrease in the abundance of proteobacterial sequences was driven by a significant decrease in the abundance of one proteobacterial class, the Deltaproteobacteria (P < 0.05) (data not shown). WWTP effluent also significantly reduced the relative abundance of Chloroflexi and Spirochaetes sequences. In contrast, there was a significant increase in the abundance of Nitrospirae sequences at downstream sites (Fig. 4). Based on SIMPER analysis, there were 17 OTUs that accounted for 20% of the variation in community composition between the upstream and downstream sites, and there were significant differences in the relative abundances of some of these OTUs (Table 3). Notable differences included significantly higher relative abundances of Sphingobacteriales, Gallionellaceae, Verrucomicrobia, and Rhodobacter sequences and significantly lower relative abundances of Crenothrix, Dechloromonas, Thiobacillus, and Desulfococcus sequences at the downstream sites (Table 3).

Fig 5 — Shannon diversity index (A) and Chao 1 richness estimator (B) based on 16S tag pyrosequencing data for sediment bacterial communities collected from rivers within two land use types (urban and suburban) from locations upstream and downstream of WWTP effluent inputs. Each data point is the mean (n = 5) ± standard error. Two-way ANOVA of Shannon data demonstrated no significant effect of land use (P = 0.328) but a significant effect of effluent input (P = 0.019) and no significant interaction effect (P = 0.936). Two-way ANOVA of Chao1 data demonstrated no significant effect of land use (P = 0.604) but a significant effect of effluent input (P = 0.000) and no significant interaction effect (P = 0.895).

Table 3.

Bacterial operational taxonomic units making the most significant contribution to variation between communities from the upstream and downstream sites

Operational taxonomic unit^c	Relative abundance (%)^a		P value^b	Contribution to variation (%)	Cumulative contribution to variation (%)	Taxonomic identification^d
Operational taxonomic unit^c	All upstream sites	All downstream sites	P value^b	Contribution to variation (%)	Cumulative contribution to variation (%)	Taxonomic identification^d
Otu4	0.16	4.35	0.043	2.51	2.51	Sphingobacteriales
Otu2	0.04	3.63	0.005	2.16	4.67	Gallionellaceae
Otu3	4.00	0.62	0.002	2.07	6.74	Crenothrix
Otu1	2.68	0.30	0.012	1.53	8.27	Dechloromonas
Otu42	0.38	2.73	0.002	1.41	9.69	Verrucomicrobia
Otu5	2.36	0.08	0.004	1.37	11.06	Thiobacillus
Otu8	2.97	0.71	<0.001	1.36	12.41	Desulfococcus
Otu39	0.10	2.10	0.272	1.22	13.64	Alteromonadaceae
Otu16	1.67	0.11	<0.001	0.94	14.58	Proteobacteria, unclassified
Otu33	0.87	2.09	0.003	0.86	15.44	Rhodobacter
Otu12	1.43	0.10	0.017	0.85	16.29	Comamonadaceae
Otu6	0.45	1.26	0.399	0.78	17.07	Deltaproteobacteria, unclassified
Otu10	0.00	1.26	0.106	0.76	17.83	Oceanospirillales
Otu7	0.03	1.23	0.239	0.73	18.55	Methylophilaceae
Otu9	0.02	1.17	0.321	0.70	19.25	Flavobacteriaceae
Otu15	0.00	1.06	0.331	0.63	19.89	Sphingobacteriales
Otu18	0.01	1.04	0.327	0.62	20.51	Methylophilus

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Each data point is the mean value (n = 5).

P value based on ANOVA comparison of all upstream and all downstream samples.

OTUs were identified within a 16S tag pyrosequencing data set based on 97% sequence identity.

Taxonomic assignments were based on comparison to the SILVA-compatible bacterial alignment database.

DISCUSSION

Effects of land use.

Freshwater ecosystems are especially susceptible to changes in land use (27), yet there is little information available on the effects of urbanization on benthic bacterial communities in lotic ecosystems (28). Comparison of the urban and suburban rivers used in this study at the reaches above the wastewater effluent inputs provides insight into the effects of land use on these communities. For example, the suburban river has agriculture in its watershed (17% of total land use), compared to no agriculture in the urban watershed, so fertilizer use may have contributed to the higher concentrations of inorganic and organic nutrients in the suburban river. Based on these higher nutrient concentrations, it is not surprising that higher numbers of heterotrophic bacteria were detected in the suburban river sediment. In contrast, the watershed of the urban river has a higher proportion of land with impervious surfaces (residential, commercial, and industrial land represented 80% of the total land use). The urban river also receives inputs from several combined sewer overflows (CSOs) that release untreated wastewater and storm water during high rainfall (www.cityofchicago.org). Non-point-source runoff from impervious surfaces and CSOs can be sources of anthropogenic pollutants, including polycyclic aromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs) (29–31), and elevated concentrations of PCBs have been reported for the NSC (32). Therefore, anthropogenic pollutants in the urban site may also have contributed to its lower numbers of heterotrophic bacteria.

Sediment bacterial communities from both rivers were dominated by Proteobacteria, a ubiquitous and metabolically diverse group of Gram-negative bacteria frequently detected in freshwater sediments (33–35). Although there was no difference in the abundance of proteobacterial sequences between the urban and suburban sites, Bacteroidetes sequences were significantly more abundant in the urban river sediment. Bacteroidetes are Gram-negative heterotrophic bacteria that are common in freshwater ecosystems and are known to degrade high-molecular-weight organic compounds (36), including petroleum hydrocarbons (37). Therefore, the higher abundance of Bacteroidetes at the urban site might have been the result of higher concentrations of complex organic compounds, including petroleum hydrocarbons (as discussed above). Another noteworthy difference in sediment bacterial communities was a 25-fold difference in the abundance of Dechloromonas sequences, which accounted for more than 5% of the total sequences in the urban sediment but less than 0.2% of the sequences in the suburban sediment. Dechloromonas species are common in aquatic sediments and are known to oxidize aromatic compounds (38), so their higher abundance in the urban sediment may reflect higher concentrations of petroleum hydrocarbons. Finally, Nitrospira sequences were 20-fold more abundant in the urban sediments. Nitrospira species are Gram-negative bacteria that catalyze the second step in the process of nitrification and are the dominant nitrite oxidizers within freshwater sediments (39). The urban sediment had a significantly higher concentration of ammonium, which represents more substrate for nitrification and could explain the higher abundance of Nitrospira sequences. In summary, significant differences in the relative abundances of several bacteria in the urban and suburban sediments, including Bacteroidetes, Dechloromonas, and Nitrospira, may be linked to anthropogenic inputs resulting from differences in land use.

Effects of effluent addition.

WWTP effluent significantly altered the downstream chemistry and bacterial communities of these rivers. In particular, the concentrations of inorganic nitrogen and phosphorus were higher downstream of the effluent, similar to what has been observed for a variety of ecosystems (4, 5, 40–42). However, it was very surprising that despite the higher concentrations of inorganic nutrients, the numbers of sediment bacteria decreased downstream of the effluent inputs. Previous studies demonstrated that increased concentrations of nitrogen and phosphorus associated with WWTPs stimulate planktonic bacterial growth (6, 7) and benthic (9) bacterial numbers. The downstream sites also had lower concentrations of sediment organic matter, which might explain the reduction in bacterial numbers. The decrease in sediment organic matter concentrations was in itself surprising, because increased concentrations of inorganic nutrients often result in greater primary production (for a review, see reference 43). However, in addition to elevated levels of nutrients, toxic compounds may also be present in WWTP effluent, and these may have inhibited bacterial populations. There is growing concern about the presence of a wide range of biologically active compounds, including antimicrobials, in rivers and streams receiving WWTP effluent in the United States (44). Many of these compounds are not completely removed by wastewater treatment, so WWTPs are point sources of these compounds (45, 46). If the effluent from the two WWTPs examined in this study contained compounds toxic to microorganisms (both algae and bacteria), this could explain both the decrease in bacterial numbers and the decrease in concentrations of sediment organic matter. Toxic compounds could also explain the observed increases in per-cell respiration rates, as previous studies indicated that respiration rates normalized by biomass increase for bacterial cells that are under stress (47). Toxic compounds in the effluent could also contribute to the decreases in bacterial diversity and species richness at the downstream sites, which conflicted with previously reported findings that demonstrated an increase in bacterial diversity downstream of a WWTP effluent input (9). Although quantification of toxic compounds in the effluents was beyond the scope of this study, future explorations of this topic are warranted.

WWTP effluent also resulted in shifts in bacterial community composition. The most striking effect was that the bacterial communities, which were clearly distinct at the upstream sites, were indistinguishable downstream of the WWTP effluents. This result provides an excellent illustration of the concept of biotic homogenization, which suggests that human modifications of the environment are reducing the biological differences that exist among natural ecosystems. As a result, there are now a series of human-altered ecosystems that consistently support a subset of naturally occurring species (48). This process is predicted to result in a more homogenized biosphere with lower diversity on regional and global scales (49). The phenomenon of biotic homogenization has been demonstrated by numerous studies focused on plant and animal communities but has been less well explored for microbial communities (48). Our results suggest that WWTP effluent may be a driver of biotic homogenization of riverine bacterial communities.

Specific changes in bacterial community composition resulted from WWTP effluent inputs. For example, there was a significant decrease in the relative abundance of proteobacterial sequences. This decrease was driven by a significant decrease in the abundance of one proteobacterial class, the Deltaproteobacteria, which includes most of the known sulfate-reducing organisms (50). For example, Desulfococcus is one sulfate-reducing genus within the Deltaproteobacteria, and the abundance of Desulfococcus sequences was significantly lower downstream of the WWTP effluent. The decreases in abundances of deltaproteobacterial and Desulfococcus sequences at the downstream sites may reflect the increased concentration of nitrate at those sites, as nitrate is a more energetically favorable electron acceptor than sulfate, and an increase in the nitrate concentration would make sulfate reducers less competitive for available electron donors.

WWTP effluent also resulted in a significant decrease in the abundance of Chloroflexi sequences. The phylum Chloroflexi includes a variety of anoxygenic phototrophic bacteria (51), so the decrease in the relative abundance of these organisms downstream of the WWTP input was somewhat surprising given the increases in concentrations of nitrogen and phosphorus. This decrease in the abundance of Chloroflexi suggests that these organisms may be sensitive to some component of the WWTP effluent. In contrast, the abundance of Nitrospirae sequences increased downstream of the WWTP effluent. Nitrospirae catalyze the second step in nitrification, so their increased abundance may be a result of the increased ammonium concentrations at the downstream sites. In addition, Nitrospirae are the dominant nitrite oxidizers within most WWTPs (52), so the observed increase in the abundance of Nitrospirae may merely be the result of a direct release of these organisms from the WWTPs. This possibility is supported by the fact that the two WWTPs included in this study do not disinfect their effluent prior to its release, which is relatively uncommon for WWTPs in the United States (53). However, previous studies indicated that most bacteria released in WWTP effluent do not typically survive in lotic systems (6), so it seems unlikely that direct inputs of Nitrospirae are the sole cause of larger populations downstream of the WWTP effluent.

Two other notable effects of the WWTP effluent were a significant decrease in the abundance of Dechloromonas sequences and a significant increase in the abundance of Sphingobacteriales sequences. Dechloromonas sequences were most abundant at the urban upstream site, and as discussed above, this may have been related to the ability of Dechloromonas to oxidize aromatic compounds and the possible presence of anthropogenic aromatic compounds (e.g., PAHs and PCBs) originating from urban runoff and combined sewer discharges. The input of WWTP effluent lowered the abundance of Dechloromonas sequences almost 10-fold, suggesting that WWTP effluent may contain lower levels of these anthropogenic aromatic compounds than untreated urban runoff and combined sewer discharges. This hypothesis is supported by previous research that demonstrated the ability of wastewater treatment processes to lower PCB and PAH concentrations in wastewater (54, 55). WWTP effluent also resulted in a higher abundance of Sphingobacteriales sequences. The Sphingobacteriales are Gram-negative bacteria that are found in a wide array of habitats and are known for their ability to utilize unusual compounds, including herbicides and antimicrobial compounds (56). The increase in the abundance of Sphingobacteriales sequences lends further support to the hypothesis that the WWTP effluent might have contained some anthropogenic compounds with antimicrobial properties.

Conclusions.

Our data demonstrate that two rivers that differed significantly in chemical and biological characteristics showed similar responses to WWTP effluent inputs, including decreases in the abundance and diversity of sediment bacterial communities, with the result that bacterial communities that were clearly distinct at the upstream sites were indistinguishable downstream of WWTP effluents. Given the ubiquity of WWTPs in the United States and worldwide, these results raise new questions about the effects of human modification of stream ecosystems. In addition, the effluent led to increased biotic homogenization, and to our knowledge, this is a new aspect of this phenomenon not previously explored. Further investigations are needed to explore the universality of biotic homogenization due to WWTP effluent across a range of river ecosystems.

ACKNOWLEDGMENTS

This work was supported by a grant to J.J.K. and E.R.-M. from the Illinois Sustainable Technology Center.

We thank David Fischer at the Cary Institute of Ecosystem Studies for analysis of dissolved organic matter. We acknowledge the technical assistance of Marty Berg and Timothy Hoellein and helpful comments on the manuscript provided by Domenic Castignetti and T. Hoellein. We also thank Pat Schloss for MOTHUR training and Peter Groffman for helpful discussions related to this work.

Footnotes

Published ahead of print 11 January 2013

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[B1] 1. US EPA. 2009. Drinking water infrastructure needs survey and assessment: fourth report to Congress, p 1–12 US EPA, Washington, DC [Google Scholar]

[B2] 2. Brooks BW, Riley TM, Taylor RD. 2006. Water quality of effluent-dominated ecosystems: ecotoxicological, hydrological, and management considerations. Hydrobiologia 556:365–379 [Google Scholar]

[B3] 3. Illinois Department of Natural Resources 2011. Illinois Coastal Management Program issue paper: Chicago River and North Shore Channel corridors. Illinois Department of Natural Resources, Springfield, IL [Google Scholar]

[B4] 4. Waiser MJ, Tumber V, Holm J. 2011. Effluent-dominated streams. Part 1: presence and effects of excess nitrogen and phosphorus in Wascana Creek, Saskatchewan, Canada. Environ. Toxicol. Chem. 30:496–507 [DOI] [PubMed] [Google Scholar]

[B5] 5. Gücker B, Brauns M, Pusch MT. 2006. Effects of wastewater treatment plant discharge on ecosystem structure and function of lowland streams. J. North Am. Benthol. Soc. 25:313–329 [Google Scholar]

[B6] 6. Garnier J, Servais P, Billen G. 1992. Bacterioplankton in the Seine River, France: impact of the Parisian urban effluent. Can. J. Microbiol. 38:56–64 [Google Scholar]

[B7] 7. Goñi-Urriza M, Capdepuy M, Raymond N, Quentin C, Caumette P. 1999. Impact of an urban effluent on the bacterial community structure in the Arga River, Spain, with special reference to culturable Gram-negative rods. Can. J. Microbiol. 45:826–832 [PubMed] [Google Scholar]

[B8] 8. Cébron A, Coci M, Garnier J, Laanbroek HJ. 2004. Denaturing gradient gel electrophoretic analysis of ammonia-oxidizing bacterial community structure in the lower Seine River: impact of Paris wastewater effluents. Appl. Environ. Microbiol. 70:6726–6737 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B9] 9. Wakelin SA, Colloff MJ, Kookana RS. 2008. Effect of wastewater treatment plant effluent on microbial function and community structure in the sediment of a freshwater stream with variable seasonal flow. Appl. Environ. Microbiol. 74:2659–2668 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B10] 10. Sander BC, Kalff J. 1993. Factors controlling bacterial production in marine and freshwater sediments. Microb. Ecol. 26:79–99 [DOI] [PubMed] [Google Scholar]

[B11] 11. Lofton DD, Hershey AE, Whalen SC. 2007. Evaluation of denitrification in an urban stream receiving wastewater effluent. Biogeochemistry 86:77–90 [Google Scholar]

[B12] 12. HDR Engineering Inc 2011. Detailed watershed plan for the North Branch of the Chicago River and Lake Michigan watershed, vol 1, prepared for Metropolitan Water Reclamation District of Greater Chicago. HDR Engineering Inc, Chicago, IL [Google Scholar]

[B13] 13. Aqua Terra Consultants 2003. Total maximum daily loads for West Branch DuPage River, Illinois. Report submitted to Illinois Environmental Protection Agency. Aqua Terra Consultants, Milwaukee, WI [Google Scholar]

[B14] 14. Findlay S, McDowell WH, Fischer D, Pace ML, Caraco N, Kaushal SS, Weathers KC. 2010. Total carbon analysis may overestimate organic carbon content of fresh waters in the presence of high dissolved inorganic carbon. Limnol. Oceanogr. Methods 8:196–201 [Google Scholar]

[B15] 15. Bear FE. 1964. Chemistry of the soil, p 70 Reinhold Publishing, New York, NY [Google Scholar]

[B16] 16. Hill B, Herlihy A, Kaufmann P. 2002. Benthic microbial respiration in Appalachian Mountain, Piedmont, and Coastal Plains streams of the eastern USA. Freshw. Biol. 47:185–194 [Google Scholar]

[B17] 17. Zuberer DA. 1994. Recovery and enumeration of viable bacteria, p 119–144 In Weaver RW, Angle S, Bottomley P, Bezdiecek D, Smith S, Tabatabai A, Wollum A, Mickelson SH, Bigham JM. (ed), Methods of soil analysis. Part 2—microbiological and biochemical properties. Soil Science Society of America, Madison, WI [Google Scholar]

[B18] 18. Kepner R, Pratt JR. 1994. Use of fluorochromes for direct enumeration of total bacteria in environmental samples: past and present. Microbiol. Rev. 58:603–615 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B19] 19. Gough HL, Stahl DA. 2003. Optimization of direct cell counting in sediment. J. Microbiol. Methods 52:39–46 [DOI] [PubMed] [Google Scholar]

[B20] 20. Boon N, Windt W, Verstraete W, Top EM. 2002. Evaluation of nested PCR-DGGE (denaturing gradient gel electrophoresis) with group-specific 16S rRNA primers for the analysis of bacterial communities from different wastewater treatment plants. FEMS Microbiol. Ecol. 39:101–112 [DOI] [PubMed] [Google Scholar]

[B21] 21. Youssef N, Sheik CS, Krumholz LR, Najar FZ, Roe BA, Elshahed MS. 2009. Comparison of species richness estimates obtained using nearly complete fragments and simulated pyrosequencing-generated fragments in 16S rRNA gene-based environmental surveys. Appl. Environ. Microbiol. 75:5227–5236 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B22] 22. Schloss PD, Westcott SL, Ryabin T, Hall JR, Hartmann M, Hollister EB, Lesniewski RA, Oakley BB, Parks DH, Robinson CJ, Sahl JW, Stres B, Thallinger GG, Van Horn DJ, Weber CF. 2009. Introducing mothur: open-source, platform-independent, community-supported software for describing and comparing microbial communities. Appl. Environ. Microbiol. 75:7537–7541 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B23] 23. Haas BJ, Gevers D, Earl AM, Feldgarden M, Ward DV, Giannoukos G, Ciulla D, Tabbaa D, Highlander SK, Sodergren E. 2011. Chimeric 16S rRNA sequence formation and detection in Sanger and 454-pyrosequenced PCR amplicons. Genome Res. 21:494–504 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B24] 24. Yue JC, Clayton MK. 2005. A similarity measure based on species proportions. Commun. Stat. Theory Methods 34:2123–2131 [Google Scholar]

[B25] 25. Shannon CE. 2001. A mathematical theory of communication. ACM SIGMOBILE Mob. Comput. Commun. Rev. 5:3–55 [Google Scholar]

[B26] 26. Chao A. 1984. Nonparametric estimation of the number of classes in a population. Scand. J. Stat. 11:265–270 [Google Scholar]

[B27] 27. Palmer MA, Covich AP, Lake S, Biro P, Brooks JJ, Cole J, Dahm C, Gibert J, Goedkoop W, Martens K, Verhoeven J, Van De Bund WJ. 2000. Linkages between aquatic sediment biota and life above sediments as potential drivers of biodiversity and ecological processes. Bioscience 50:1062–1075 [Google Scholar]

[B28] 28. Perryman SE, Rees GN, Walsh CJ. 2008. Analysis of denitrifying communities in streams from an urban and non-urban catchment. Aquat. Ecol. 42:95–101 [Google Scholar]

[B29] 29. Lau SL, Han Y, Kang JH, Kayhanian M, Stenstrom MK. 2009. Characteristics of highway stormwater runoff in Los Angeles: metals and polycyclic aromatic hydrocarbons. Water Environ. Res. 81:308–318 [DOI] [PubMed] [Google Scholar]

[B30] 30. Walker WJ, McNutt RP, Maslanka CAK. 1999. The potential contribution of urban runoff to surface sediments of the Passaic River: sources and chemical characteristics. Chemosphere 38:363–377 [DOI] [PubMed] [Google Scholar]

[B31] 31. Polls I, Lue-Hing C, Zenz DR, Sedita SJ. 1980. Effects of urban runoff and treated municipal wastewater on a man-made channel in northeastern Illinois. Water Res. 14:207–215 [Google Scholar]

[B32] 32. Illinois EPA. 2004. Illinois water quality report IEPA/BOW/04–006. Illinois EPA, Springfield, IL [Google Scholar]

[B33] 33. Tamaki H, Sekiguchi Y, Hanada S, Nakamura K, Nomura N, Matsumura M, Kamagata Y. 2005. Comparative analysis of bacterial diversity in freshwater sediment of a shallow eutrophic lake by molecular and improved cultivation-based techniques. Appl. Environ. Microbiol. 71:2162–2169 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B34] 34. Methe BA, Hiorns WD, Zehr JP. 1998. Contrasts between marine and freshwater bacterial community composition: analyses of communities in Lake George and six other Adirondack lakes. Limnol. Oceanogr. 43:368–374 [Google Scholar]

[B35] 35. Fazi S, Amalfitano S, Pernthaler J, Puddu A. 2005. Bacterial communities associated with benthic organic matter in headwater stream microhabitats. Environ. Microbiol. 7:1633–1640 [DOI] [PubMed] [Google Scholar]

[B36] 36. Thomas F, Hehemann JH, Rebuffet E, Czjzek M, Michel G. 2011. Environmental and gut bacteroidetes: the food connection. Front. Microbiol. 2:93 doi:10.3389/fmicb.2011.00093 [DOI] [PMC free article] [PubMed] [Google Scholar]

[B37] 37. Zhang DC, Mörtelmaier C, Margesin R. 2012. Characterization of the bacterial archaeal diversity in hydrocarbon-contaminated soil. Sci. Total Environ. 421-422:184–196 [DOI] [PubMed] [Google Scholar]

[B38] 38. Coates JD, Chakraborty R, Lack JG, O'Connor SM, Cole KA, Bender KS, Achenbach LA. 1980. Anaerobic benzene oxidation coupled to nitrate reduction in pure culture by two strains of Dechloromonas. J. Volcanol. Geotherm. Res. 7:443–481 [DOI] [PubMed] [Google Scholar]

[B39] 39. Altmann D, Stief P, Amann R, De Beer D, Schramm A. 2003. In situ distribution and activity of nitrifying bacteria in freshwater sediment. Environ. Microbiol. 5:798–803 [DOI] [PubMed] [Google Scholar]

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PERMALINK

Wastewater Treatment Effluent Reduces the Abundance and Diversity of Benthic Bacterial Communities in Urban and Suburban Rivers

Bradley Drury

Emma Rosi-Marshall

John J Kelly

Abstract

INTRODUCTION