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a Laboratory of Soil Ecology and Microbiology, University of Rhode Island, Kingston, RI 02881
b Geomatrix LLC, Killingworth, CT 06419
c Department of Crop, Soil, and Environmental Sciences, University of Arkansas, Fayetteville, AR 72701
* Corresponding author (jamador{at}uri.edu)
Received for publication October 14, 2005.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Abbreviations: AIR, aerated lysimeters • BOD5, 5-d biological oxygen demand • CLPP, community-level physiological profile • DGGE, denaturing gradient gel electrophoresis • LEACH, lysimeters vented to leachfield • MPN, most-probable number • PCR, polymerase chain reaction • PLFA, phospholipid fatty acid • SRB, sulfate-reducing bacteria
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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The ability of soil absorption fields to improve water quality depends largely on the types of organisms present and their relative abundance (structure) and the processes they carry out (function). Organic C removal and nutrient transformations are biochemical processes, primarily performed by microorganisms, whereas soil fauna feeding on microorganisms are thought to control pathogens. The leachfield soil community develops its structure and function from: (i) the organisms originally present in the soil, (ii) colonization of the soil by organisms present in septic tank effluent, (iii) the type and amounts of electron donors and acceptors, (iv) the physical and chemical conditions (e.g., pore size, pH, temperature), and (v) the interactions among members of the food web.
Current understanding of the microbial ecology of leachfield soils, conventional or otherwise, is sparse. A number of studies have examined the presence of specific groups of microorganisms (Calaway et al., 1952; Pell et al., 1990) and of meso- and macrofauna (Brink, 1969) in absorption fields. In terms of function, the emphasis has been on ecological interactions that remove pathogens (Bomo et al., 2004) and affect microbial population dynamics (Woombs and Laybourn-Parry, 1986, 1987), and on nutrient transformations (Walker et al., 1973; Kristiansen, 1981a).
In a previous study we found that active aeration of model leachfields improved the removal of N, fecal coliforms, and BOD5 (5-d biological oxygen demand) relative to conventional (unaerated) mesocosms (Potts et al., 2004). The intermittent introduction of air appeared to overcome limitations on nitrification, subsequent denitrification, and oxidation of biodegradable C, while enhancing pathogen removal. These improvements were observed in the absence of a clogging layer or biomat. Aeration appeared to cause a major shift in the function of the soil community that improved water quality and reduced the risk of hydraulic failure by clogging. This change in function is probably the result of differences in faunal and microbial community structure.
Because the leachfield aeration technology employed in our study (Potts et al., 2004) is a recent innovation (Potts, 2000), no data are available on its effects on the microbial ecology of leachfield soil. In this study, we characterized the microbivorous fauna (protozoa and nematodes) and microflora associated with soils in the model leachfields under aerated and unaerated (conventional) conditions employed in Potts et al. (2004). Microbial community structure was examined using both culture-independent (PLFA [phospholipid fatty acid] analysis [Guckert et al., 1985], PCR-DGGE [polymerase chain reaction–denaturing gradient gel electrophoresis; Amann et al., 1995]) and culture-dependent (Biolog EcoPlates) methods, where the latter is more closely linked to function than the former because it measures substrate use. Protozoa and nematodes were characterized using MPN (most-probable number; Ingham, 1994a) and direct enumeration (Ingham, 1994b) techniques, respectively. Analyses were conducted on soil from different depths because soil communities are expected to stratify in systems with unidirectional inputs, such as leachfield soils. The analyses were conducted in aerated (AIR) and unaerated (LEACH) mesocosms to examine the differential effects of the presence of O2 on the community ecology of leachfield soils. Silica sand was used in our experiment because it is relatively inert and presumed to be devoid of microorganisms, allowing evaluation of the effects of aeration on water quality under more stringent conditions than expected using native soil. Because aeration removes the energetic constraints imposed by the low levels of O2 found in leachfield soil, we hypothesized that the faunal and microbial communities in aerated leachfield soil would be larger and more diverse than in unaerated soil.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Treatments
A piston pump was used to introduce ambient air at regular intervals into the headspace of AIR lysimeters to maintain an O2 level of 0.20 to 0.21 mol mol–1. This resulted in a positive pressure (2.5–6.7 kPa) within AIR lysimeters. To mimic the in situ composition of the atmosphere found in a conventional leachfield, the headspace and the gravel bed below the soil of LEACH lysimeters were vented to a septic system leachfield trench in which the O2 concentration ranged from <0.021 to 0.16 mol mol–1 (Potts et al., 2004). Each treatment was replicated three times.
Soil Sampling and Processing
Soil cores were obtained on 25 June 2003 from the top 13 cm of each lysimeter using a cut-off, 60-mL plastic syringe (2.54 cm i.d., 13.4 cm long) that was surface sterilized with a 70% methanol solution before sampling. Wastewater remaining on the surface of the soil in the mesocosms was siphoned off before sampling. Three cores were removed from each mesocosm within a 5-cm radius. Cores were cut into 0- to 4- and 4- to 13-cm sections, placed in sterile plastic bags, and kept on ice during transport to the laboratory in Kingston, RI. Equal amounts of soil from a single lysimeter were mixed to produce one composite sample per lysimeter for each depth. One composite soil sample per lysimeter and depth was employed for determination of CLPPs (community-level physiological profiles), enumeration of protozoa and nematodes, enumeration of Escherichia coli, and analysis of community structure by PCR-DGGE. Analysis of community structure using PLFA was conducted on a single composite sample from all three lysimeters within a treatment. Samples for analysis by PLFA and PCR-DGGE were kept at –80°C and shipped by overnight courier within 24 h of collection to Knoxville, TN, and Fayetteville, AR, respectively. Samples for analysis of protozoa and nematodes were kept at 4°C and shipped to Corvallis, OR, by overnight courier within 24 h of collection.
Faunal Community Analysis
Enumeration and identification of protozoa and nematodes were conducted by Soil FoodWeb (Corvallis, OR). Protozoa were enumerated using a MPN technique (Ingham, 1994a), a method based on determination of the presence or absence of microorganisms in individual portions of several consecutive dilutions of soil (Alexander, 1982), and were grouped into flagellates, ciliates, and amoebae. Nematodes were extracted using the sugar flotation method (Ingham, 1994b) and enumerated and assigned to trophic groups based on microscopic examination. Analyses were conducted on one composite sample from each replicate lysimeter.
Phospholipid Fatty Acid Analysis
The PLFA analysis was performed by Microbial Insights (Rockford, TN). Composite samples from three replicate lysimeters per treatment and depth were used for analyses. The PLFAs were extracted using a modification (White et al., 1979) of the method of Bligh and Dyer (1959). Fatty acid methyl esters were separated by gas chromatography and identified by retention time and mass spectrometry as described by Tunlid et al. (1989). The detection limit was 7 pmol PLFA. For the purposes of community structure analyses, PLFAs were divided into markers for six different groups (Dowling et al., 1986; Edlund et al., 1985; White et al., 1996, 1997): anaerobic Gram-negative bacteria/firmicutes (i14:0, i15:0, 
15:0, i16:0, i17:0, 17: 0), proteobacteria (16:1
9c, 16:17c, 16:17t, 16:15c, 17:16c, cy17:0, 18:19c, 18:17c, 18:17t, 18:15c, 19:1, 19:1ß, cy19:0), anaerobic metal reducers (i15:1, i15:1ß, i16:1, i17:17c, br19: 1), actinomycetes/sulfate-reducing bacteria (br16:0, 10me16:0, 12me16:0, 10me17:0, 11me17:0, 12me17:0/18:2, 10me18:0, 12me18:0), general (14:0, 15:0, 16:0, 17:0, 18:0, 20:0, 22:0), and eukaryotes (18:2, 18:2ß, 18:26, 18:33, 20:46, 20:53). The relative contribution of each group to the total PLFA pool was determined from the sum of the amount of all PLFA markers for a particular group divided by the total amount of PLFA in the sample. The biomass ratio of prokaryotes to eukaryotes was determined from the sum of PLFA markers for firmicutes, proteobacteria, anaerobic metal reducers, actinomycetes, and general bacteria divided by the eukaryotic PLFA. The starvation index was determined from the sum of the ratios of cy17:0 to 16:17c and cy19:0 to18:17c (Kieft et al., 1994). The membrane stress index was calculated from the sum of 16:17t/16:17c and 18:17t/18:17c (Guckert et al., 1985, 1986). A factor of 20000 cells pmol–1 of PLFA—developed by Microbial Insights—was used to calculate the number of viable organisms.
Community-Level Physiological Profiles
EcoPlates (Biolog, Hayward, CA) were used to determine CLPPs using the method of Gamo and Shoji (1999), which involves determining the MPN of bacteria capable of growing on 31 different substrates. Freshly sampled soil from both depths (1.0 g) was diluted 1:10 with 50 µM phosphate buffer (pH 7.2). The resulting suspension was shaken vigorously in a reciprocal shaker for 2 h (Balser et al., 2002) and the soil allowed to settle before sampling the supernatant liquid. Six 10-fold dilutions were used, with each dilution replicated three times. Well inoculation volume was 150 µL. After inoculation, plates were incubated at 22°C for 6 d. Optical density (OD) was measured at 595 nm with a Model EL311 Automated Microplate Reader (Bio-Tek Instruments, Winooski, VT). The minimum OD for positive wells was 0.30. The MPNs were determined as described by Woomer (1994). Replicates from each treatment/depth combination were analyzed independently.
Polymerase Chain Reaction–Denaturing Gradient Gel Electrophoresis Analysis
The DNA was extracted from soil following the protocol for the bead-beating FastDNA Spin kit for soil (Qbiogene, Carlsbad, CA) and quantified by labeling with Hoeschst 33258 dye and measuring fluorescence, using calf thymus DNA as a standard. Extracted DNA was used as a template and amplified by PCR with primers 338FGC and 518R (Övreås et al., 1997), which are specific for 16S rRNA gene fragments of eubacteria. The PCR conditions for each 100-µL reaction were: 10 mM Tris–HCl, 50 mM KCl, 0.01% (w/v) gelatin, 1.5 mM MgCl2, 0.2 µM each primer, 200 µM each deoxynucleotide, 400 ng µL–1 bovine serum albumin, 2.5u of Taq DNA polymerase (Promega, Madison, WI), and 2 µL of template DNA. The PCR was performed in a MJ Research thermal cycler (Waltham, MA) with the following conditions: initial denaturation at 94°C for 5 min, followed by 30 cycles of 94°C for 30 s, 55°C for 30 s, and 72°C for 30 s, and a final extension at 72°C for 20 min. The PCR products were resolved on 1.0% agarose gels stained with ethidium bromide, visualized using a Kodak EDAS 290 system, and analyzed using Kodak 1D software (Kodak, New Haven, CT). The DNA content in agarose gels was quantified by comparison with a precision molecular mass standard. Approximately 1800 ng of PCR product was loaded onto polyacrylamide DGGE gels for generation of community profiles. Electrophoresis was run in a D-Code system (Bio-Rad Laboratories, Hercules, CA) on a 1.5-mm-thick vertical gel containing 8% polyacrylamide (37.5:1 acrylamide/bisacrylamide) with a linear gradient of 55 to 70% denaturant solutions (where 100% is equivalent to 7 M urea and 40% deionized formamide). Gels were run in TAE buffer (40 mM Tris, 20 mM acetic acid, and 1 mM EDTA [ethylenediaminetetraacetic acid] at pH 7.6) for 16 h at 70 V and 60°C and stained for 30 min in SYBR Green I dye. Gels were pictured using a Kodak EDAS 290 system and analyzed using Quantity One software (Bio-Rad). All samples were run on a single DGGE gel to enable the software to detect and match bands based on migration distances.
Enumeration of Escherichia coli
The number of E. coli in freshly sampled soil was determined using the method described by Turco (1994).
Data Analysis
Differences between treatments in the numbers of protozoa, nematodes, and microorganisms growing in the CLPP assay, E. coli, and DNA concentration were determined using Student's t-test. A one-way analysis of variance and a pairwise multiple comparison procedure were used to determine differences in diversity indices.
The DGGE and PLFA data were analyzed by constructing a similarity matrix using binary data (presence and absence of bands only) and the Dice similarity coefficient (Cs) was calculated using the equation:
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RESULTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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A total of 45 different PLFAs were detected in all the soil samples, of which 30 were common to all samples. The total number of detectable PLFA markers—a measure of species richness—extracted from AIR soil was higher than for LEACH soil (Table 3). Other diversity indices (H and E) were similar for both treatments (Table 3). Values of Cs indicated that the microbial communities of soils in the LEACH treatment at 0 to 4 and 4 to 13 cm were most similar (0.94), followed by AIR and LEACH soil at 4 to 13 cm (0.90), AIR soil at 0 to 4 and 4 to 13 cm (0.89), with AIR and LEACH soil at 0 to 4 cm being least similar (0.87).
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and 19:1ß) were present exclusively in AIR soil, with a third marker, 17:16c, present only in soil from 0 to 4 cm in AIR and LEACH soil (Table 4). Of the markers for anaerobic metal reducers, i15:1 was absent from LEACH surface soil, and i15:1ß was absent from LEACH soil at both depths. Six markers for actinomycetes/SRB were absent from soil at both depths in the LEACH treatment, whereas one and three markers were absent at 0 to 4 and 4 to 13 cm in AIR soil, respectively. Of the general markers for microorganisms, 22:0 was present only in surface AIR soil. Of the six eukaryotic markers detected, 18:33 was absent from AIR soil at both depths. In LEACH soil, 20:53 was absent at both depths, and 18:2 was absent at 4 to 13 cm.
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-glycerol phosphate, L-phenylalanine, L-threonine, glycyl-L-glutamic acid, and phenylethylamine (Fig. 2). The LEACH soil had higher numbers of microorganisms that grew on -D-lactose and D-malic acid. Soil at 4 to 13 cm had values similar to those for surface soil, although numbers of organisms were lower for both treatments (data not shown).
Numbers of Escherichia coli
Numbers of E. coli in AIR and LEACH in surface soil were 5 x 103 and 6.2 x 104 CFU kg–1, respectively. At 4 to 13 cm, AIR and LEACH soil had 6 x 103 and 5.0 x 104 CFU E. coli kg–1, respectively. Differences between treatments at a particular depth were statistically significant.
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Conditions in the LEACH soil lysimeters had a negative effect on microbivorous fauna. Numbers of protozoa in the flagellate and amoebae groups were one to three orders of magnitude lower in LEACH soil, and ciliates were absent from this treatment. Furthermore, nematodes were absent from LEACH soil. The smaller contribution of eukaryotic PLFAs to the LEACH microbial community supports the data obtained by enumeration of protozoa and nematodes. Although some species of free-living nematodes are capable of surviving periods of anoxia (Wright, 1998), and some parasitic species remain viable after prolonged exposure to anaerobic conditions (Stott, 2003), most protozoa and bacterivorous nematodes use O2 as a terminal electron acceptor, a condition favored in the AIR treatment.
Grazing by micro- and mesofauna is often cited as a mechanism of removal for pathogenic organisms in soil infiltration systems (Weber-Shirk and Dick, 1997). If this is the case, this mechanism is more likely to be of significance in AIR soil. Enumeration of E. coli in soil showed that LEACH soil contained a significantly higher number of these bacteria than AIR soil. Previous results (Potts et al., 2004) showed that the number of fecal coliform bacteria in drainage water was also significantly lower in the AIR treatment. Bomo et al. (2004) observed that removal of Aeromonas hydrophila, a fish pathogenic bacterium, in sand filters was proportional to the number of protozoa present in sand.
Consumption of bacteria by protozoa and nematodes can enhance bacterial metabolism (Griffiths, 1994; Woombs and Laybourn-Parry, 1986), the release of mineral N from bacteria (Amador and Görres, 2004), and the activities of soil nitrifiers (Griffiths, 1989). As such, microbivorous grazing may have important consequences for C and N use in AIR and LEACH soil. Higher numbers of microbivorous fauna in AIR soil, and consequent grazing of microorganisms, requires additional C consumption by microorganisms to maintain the microbial community at steady state. This greater need for C may be reflected in the complete consumption of BOD5 in AIR lysimeters (Potts et al., 2004) and would contribute to the high starvation index exhibited by the microbial community in AIR soil. With respect to N, grazing by microbivorous fauna results in the mineralization and excretion of a higher proportion of bacterial N because of the disparity in C/N ratios of the grazing fauna (5–6) and bacteria (4–5) (Amador and Görres, 2004). Nitrogen is excreted as NH4 (Griffiths, 1994; Wright, 1998), which can be oxidized to NO3. Enhanced N mineralization and subsequent nitrification would support higher rates of denitrification and greater N removal in AIR lysimeters, as observed previously (Potts et al., 2004).
Structure and Function of the Microbial Community
As was the case for the microbivorous fauna, the microbial community that developed during 26 mo in the mesocosms probably had its origins in wastewater inputs, inasmuch as the sand medium used was of synthetic origin.
The total number of microorganisms calculated based on PLFA in AIR soil is of the same order of magnitude (1012 cells kg–1) as that observed by Kristiansen (1981b) in sand filter trenches at comparable temperatures and wastewater loads. Differences in microbial biomass as a result of aeration and depth are supported by differences in the concentration of DNA extracted from soil and by the average number of microorganisms per substrate detected in the CLPP assay. Lower microbial biomass in LEACH soil may be responsible for incomplete use of BOD5 (Potts et al., 2004), and probably reflects electron acceptor limitations. Microbial biomass decreased with depth in both treatments, and was lower in LEACH than in AIR treatment. Lower microbial biomass with depth is typical of surface soils, where inputs of matter and energy are unidirectional, and has been observed for numbers of aerobic bacteria in soil absorption systems (Pell and Nyberg, 1989; Pell et al., 1990). Differences in biomass between AIR and LEACH treatments would be expected as a result of the differential availability of electron acceptors (Fuhrmann, 2004).
The AIR soil had greater microbial diversity, as indicated by a higher species richness based on PLFA markers and 16S rRNA gene fragments in AIR than in LEACH soil at both depths. In addition to fewer total DNA fragments from LEACH soil that separated as distinct bands using DGGE, the phylotypes amplified most strongly were different between the LEACH and AIR treatments (data not shown). Furthermore, the separation of LEACH and AIR soils by cluster analyses suggests that, despite common phylotypes, there are distinct differences in the dominant community members between these treatments.
The bulk of the microbial community in both treatments consisted of proteobacteria, a group that uses a wide variety of C sources, adapts quickly to changes in environmental conditions, and dominates the microbial community in activated sludge (Snaidr et al., 1997; Wagner et al., 1993). It includes enteric bacteria, Pseudomonas species, and chemolithotrophs. The AIR soil had the largest proportion of PLFA markers within the actinomycete/SRB group. Actinomycetes require oxic conditions, prefer a near-neutral pH, and are capable of using a number of the polysaccharides, proteins, and fats found in domestic wastewater. As such, they would be expected to thrive in AIR leachfield soil. Some midchain branched saturated PLFAs found in actinomycetes are also associated with SRB, such as Desulfobacter (Dowling et al., 1986), which should be found preferentially in LEACH soil. This was not the case, suggesting that the difference between AIR and LEACH treatments is attributable to actinomycetes. Kristiansen (1981b) found that streptomycetes were present in sand-filter trenches, although their numbers did not change in response to loading with septic tank effluent.
The contribution of anaerobic Gram-negative bacteria/firmicutes to the community in LEACH soil was roughly twice that in AIR soil. These organisms can be of fecal origin, with members ranging from strictly aerobic to anaerobic. Their contribution to the LEACH soil microbial community probably reflects the lower O2 concentration in this treatment.
A number of factors probably contribute to lower diversity in LEACH soil. Ecosystems in which physicochemical factors dominate tend to have lower microbial diversity (Atlas and Bartha, 1998). In our study, the persistent low levels of O2 limit the microbial community to those that prefer or can tolerate these conditions, including microaerophilic, obligate anaerobic, and facultative anaerobic organisms. In addition, the presence of elevated concentrations of H2S (Potts et al., 2004)—which is toxic to many aerobic microorganisms (Postgate, 1984)—further restricts the community to those microorganisms capable of either detoxifying or otherwise tolerating high concentrations of this gas. A higher membrane stress index in LEACH than AIR soil suggests that toxicity may be involved in determining the structure of the LEACH microbial community. Differences in the size of the population of the microbivorous fauna may also have a differential effect on microbial community composition. The microbial community structure of soil changes in response to the presence of microbivorous nematodes (Griffiths et al., 1999) and protozoa (Rønn et al., 2002). Selective grazing by protozoa also affects the composition of the microbial community in activated sludge (Güde, 1979). Soil pH, which can exert selective pressure on the microbial community, ranged from 5.9 in AIR soil to 6.4 in LEACH soil, suggesting that it was not an important determinant of microbial community structure.
In contrast to the results of the culture-independent assays, CLPP analysis of the microbial community found little difference in functional diversity between AIR and LEACH soil. The reason for this discrepancy probably lies in the subset of the microbial community that was evaluated by these different methods. As employed in this study, the EcoPlate–MPN method would detect only those microorganisms capable of growing under the aerobic conditions of the assay—that is, aerobes and facultative anaerobes. The wide range of electron acceptors used by these microorganisms makes them more likely to be found in both AIR and LEACH soil under the conditions of the assay.
The difference in starvation index between AIR and LEACH soil is puzzling, given the considerable amount of organic C present in septic tank effluent. A number of factors can lead to an elevated starvation index. For example, the cyclopropyl content of phospholipids fatty acids increases during the stationary growth phase (Kieft et al., 1994). Other conditions known to result in elevated cyclopropyl levels include low C concentrations, high acidity, low O2 levels, and high temperature, as observed in pure cultures of E. coli by Knivett and Cullen (1965). Given the high concentration of C in wastewater applied equally to both treatments, high levels of dissolved O2 found in AIR drainage water (Potts et al., 2004), the relatively mild temperatures under which the experiment was conducted, and small differences in soil pH, other explanations must be considered.
The paradox of "starvation in the midst of plenty" may be explained by differences in microbial growth phase. Assuming the lysimeters are at steady state with respect to the size of the microbial population (a reasonable assumption after 2 mo under the same conditions [Pell and Nyberg, 1989]), the microbial population in AIR soil may be kept in the exponential growth phase by the frequent input of substrates. Under these circumstances, the size of the population would be controlled by the amount of C input into the system. The complete absence of BOD5 in the effluent of AIR lysimeters (Potts et al., 2004) indicates that C availability limits microbial growth, supporting the contention that the microbial population is in the exponential growth phase. The grazing pressure exerted by larger numbers of microbivorous fauna in AIR soil may further contribute to the starvation phenomenon by keeping a fraction of the microbial population in a constant state of growth. In contrast, the relatively high BOD5 in the drainage water from LEACH lysimeters (Potts et al., 2004) suggests that factors other than C availability, such as availability of electron acceptors or toxicity of H2S, limit the size of the population. Lower grazing pressure by microbivorous fauna may further contribute to the differences between AIR and LEACH treatments.
The controlled conditions of our experiment allowed us to evaluate the role of aeration in shaping the faunal and microbial communities of leachfield soils. The various spatial and temporal heterogeneities in wastewater and in soil physical, chemical, and biological properties and processes encountered under field conditions undoubtedly also affect the structure and function of these communities. This is particularly likely in shallow trench leachfields, where the abundance of autochthonous microflora and fauna is high and temperature variations are not as dampened and uniform as in conventional fields with greater depth. Our results indicate that the response of the faunal and microbial communities to aeration warrants further examination in the field. Establishment of clear links between the leachfield soil community and the biochemical processes involved in wastewater treatment will advance our understanding of how these systems function and enhance our ability to optimize their water quality functions.
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ACKNOWLEDGMENTS |
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