Ecological Engineering 23 (2004) 107–115

Plant effects on microbial assemblages andremediation of acidic coal pile runoff in

mesocosm treatment wetlands

Beverly Collins∗, J. Vaun McArthur, Rebecca R. Sharitz

Savannah River Ecology Laboratory, P.O. Drawer E, Aiken, SC 29802, USA

Received 16 December 2003; received in revised form 8 July 2004; accepted 14 July 2004

Abstract

Constructed treatment mesocosm wetland systems comprised of an anaerobic wetland followed by two aerobic wetlands anddiffering in aerobic wetland composition (real plants, plastic plants, no plants) and order (shallow aerobic wetland followed bydeep aerobic wetland or vice versa) were compared to determine if composition or order affect remediation of acidic, metalcontaminated water associated with a coal-fired power plant. Aerobic cell composition was varied to determine if native plantshave no function, are only structure for bacteria, or improve water quality. The anaerobic cells increased pH from 2.4 to 6.4 andreduced Al, Fe, and Zn by 93–99%. Order of the aerobic cells did not affect water quality. Cells with real versus plastic plants haddifferent bacterial assemblages, but these were not clearly related to water quality differences between the tank types. Cells withreal plants had lower pH and NH4 and higher Fe and Mn, especially during the growing season, than those with plastic plantso s. Planteda e and MncP

K

1

d

f

d totna-

001;inub-andssul-

0

r without plants. We conclude that plants do affect water quality, in part because they affect bacterial assemblageerobic wetlands can help remediate acidic, metal-contaminated water, but may contribute to lower pH and greater Foncentrations in treatment effluent.ublished by Elsevier B.V.

eywords:Coal pile runoff basin; Constructed treatment wetlands; Acidic drainage

. Introduction

Passive treatment systems such as anoxic limestonerains (ALDs), successive alkalinity producing sys-

∗ Corresponding author. Tel.: +1 803 725 8158;ax: +1 803 725 3309.

E-mail address:collins@srel.edu (B. Collins).

tems (SAPs), and constructed wetlands are useremediate contaminated mine drainage (Wildeman eal., 1993; Hedin et al., 1994; Barton and Karathasis, 1999; Brenner, 2001; Mays and Edwards, 2Ye et al., 2001). For example, anaerobic wetlandswhich water flows through limestone and organic sstrate can treat acidic coal drainage. These wetlgenerate alkalinity through bacterially mediated

925-8574/$ – see front matter Published by Elsevier B.V.doi:10.1016/j.ecoleng.2004.07.005

108 B. Collins et al. / Ecological Engineering 23 (2004) 107–115

fate reduction and limestone dissolution and raise pH,which can increase adsorption and coprecipitation ofmetals (Wildeman et al., 1993; Williams and Stark,1997; Mitsch and Wise, 1998; Barton and Karathana-sis, 1999; Thomas et al., 1999; Lee et al., 2002). Aer-obic wetlands, often planted with emergent or sub-merged vegetation, can treat net alkaline drainage orserve as polishing systems (Wildeman et al., 1993; Sk-ousen et al., 1998; Brenner, 2001; Mays and Edwards,2001).

Treatment effectiveness varies among passive sys-tems (Mays and Edwards, 2001) and can be influencedby system design. For example, the effectiveness ofaerobic wetlands likely depends on the presence andspecies composition of vegetation; microbial compo-sition and activity; and wetland physical characteristicssuch as depth.

Vegetation in wetlands can contribute to remedi-ation by accumulating contaminants, influencing themicrobial community, or functioning as sites of pre-cipitation and sedimentation (Wildeman et al., 1993;Gazea et al., 1996; Zayed et al., 1999; Mays and Ed-wards, 2001; Ye et al., 2001). Several native wetlandplants have been shown to accumulate metals (Zayedet al., 1999), although the total metal content of plantsusually accounts for only a small percentage of themetal load in the wetland (Mays and Edwards, 2001;Ye et al., 2001). In addition, plants provide substrateand can modify the local root and stem environment(Gazea et al., 1996); both of these attributes can affectm in-fl theb nof vew

wa-t fec-t atedt sh cies,w aveg mer-g ,1 rd cro-b

nd-i sys-

tems that have different combinations of functionalattributes. We compared the effectiveness of meso-cosm systems containing deep (60 cm water depth) andshallow (20 cm water depth) aerobic wetlands that dif-fered in plant species composition and arrangement ofthe deep and shallow wetlands (Fig. 1). Each systemcontained an anaerobic wetland connected in series tothe two aerobic wetlands. The anaerobic wetland in-creased pH (from 2.4 to 6.4) and removed 93–99% ofAl and Fe in acidic, metal contaminated runoff from acoal pile runoff basin (CPRB) (Table 1; Thomas et al.,1999; Thomas, 2003). The aerobic wetlands were de-signed to further remediate the less acidic water. Theywere (a) left unplanted, (b) planted with plastic plants,or (c) planted with native species (emergent speciesJuncus effusesandPontedaria cordatain the shallowwetlands, submerged speciesMyriophyllumaquaticumand floating-leaved speciesNymphaea odoratain thedeep wetlands) to determine if plants have no effect,serve only as structure for microbes, or contribute toremediation. To determine if depth of the wetland thatreceives contaminated water influences treatment ef-fectiveness, we varied the order of the deep and shallowwetlands in the treatment system.

2. Materials and methods

2.1. Constructed treatment wetland system (CTW)design

on-t ento en,S ma creC tra-tF(

u-o ms,wo A( ov-e vol-u om-p ged

icrobial species composition and activity and canuence treatment effectiveness if microbes provideulk of treatment activity. Thus, plants may haveunction, are only structure for bacteria, or improater quality.Physical characteristics of wetlands, such as

er depth at the inlet, could influence treatment efiveness. Metal concentration in plants often is relo rooting depth (Guilizzoni, 1991). Deeper wetlandarbor submerged and floating-leaved plant spehich are usually shallow-rooted and tend to hreater metal concentrations than deep-rooted eent species of shallow wetlands (Sparling and Lowe997; Mays and Edwards, 2001). In addition, wateepth can influence redox conditions and affect miial assemblages.

One approach to decreasing variability and fing effective treatment systems is to compare

The array of CTWs received acidic, metal caminated effluent from a CPRB on the Departmf Energy’s Savannah River Site (SRS) near AikC, USA. The CPRB receives 169 l/min runoff frocoal-fired power plant coal pile. The 12.5 a

PRB has low pH (2.0–3.0) and high concenions of SO4

2− (1.1–3.3 g/l), Al (66.7–70.5 mg l−1),e (114–152 mg l−1), and Mn (4.74–5.00 mg l−1)

WSRC, 2000).The CTW array was a noncirculating, contin

us flow system of 48 mesocosm wetland systeith three wetlands (A–C) per system (Fig. 1). Aer-bic CPRB effluent flowed into anaerobic wetland0.73 m2 surface area by 1.2 m deep), which was cred to prevent evaporation and contained (85% byme) a mixture (1:3) of limestone screenings and costed stable waste. Flow rates into wetland A ran

B. Collins et al. / Ecological Engineering 23 (2004) 107–115 109

Fig. 1. One of the mesocosm treatment wetland systems. Water flows from the CPRB through anaerobic wetland A, aerobic deep or shallowwetland B, aerobic deep or shallow wetland C, then back to the CPRB.

between 16 and 28 ml/min and averaged 20 ml/min(Thomas, 2003). The effluent from A flowed througha series of two aerobic wetlands (B and C), one deep(0.73 m2 surface area by 1.2 m deep) and one shallow(0.73 m2 surface area by 0.8 m deep), before returningto the CPRB (Fig. 1). Deep aerobic wetlands contained,by volume, 50% fine-grain sand and had 0.6 m waterdepth; shallow wetlands contained 75% fine-grain sandand had 0.2 m water depth. Water depths of the deepand shallow wetlands were based on rooting depths ofthe plant species and depths of natural wetlands wherethese species occur.

Aerobic wetlands differed among the systems in at-tributes that were hypothesized to influence treatmenteffectiveness. To determine if the order of deep andshallow aerobic wetlands influences treatment effec-tiveness, half the systems (24) had a deep wetland Band a shallow wetland C; the other half had a shallowwetland B and a deep wetland C. To determine if plantsenhance water treatment either by concentrating metalsor providing substrate for microbes, 16 randomly cho-sen systems had native aquatic macrophytes planted inthe aerobic tanks, 16 systems contained no plants; and16 had artificial plants of each “species”.

Plants for the CTW were collected during the dor-mant season, between January and February, 1999,from ephemeral ponds and reservoirs on the SRS thatdo not receive coal runoff. Emergent species,JuncuseffususandP. cordata, were planted in the shallow aer-obic wetlands; a floating-leaved species,N. odorata,a di areaw vid-u lastic

plants mimicking the growth forms and geometry ofNymphaeaandJuncuswere placed in the wetlands thatcontained artificial plants. These plastic plants couldserve as substrate for bacteria, but lacked the biochem-ical interactions of real plants; we did not attempt to‘grow’ the plastic plants as the real plants grew overtime.

2.2. Water sampling and analyses

To determine the treatment effectiveness of the aer-obic wetland combinations, water samples, as well astemperature, and pH were taken monthly between April1999 and March 2000 from the effluent of row C aer-obic tanks, and from April to September 2000 fromthe effluent of both aerobic tanks. Temperature and pHwere measured with a calibrated Model 250A OrionpH meter. Water samples were filtered and analyzedfor DOC by a Shimadzu Total Organic Carbon Ana-lyzer (TOC-5000A); NO3-N, NH4-N, and orthophos-phate by a Technicon Auto-Analyzer II, Method 696-82W; metals by a Perkin Elmer ICP-MS and sulfateby a Dionex DX500 Ion Chromatograph, EPA Method300.

2.3. Microbial community analysis

Plant samples (both plastic mimics and real plants)were aseptically collected by clipping with sterile scis-sors and forceps. For theNymphaeasamples, one leafa ol-l fs . ForJ

nd a submerged species,M. aquaticum, were planten the deep aerobic wetlands. Each wetland surfaceas divided into five equal sections, and one indial of each species was planted in each section. P

nd 10 cm of petiole per individual plant were cected, and forMyriophyllum, one 10 cm portion otem with leaves attached, was collected per plantuncus, one stem per plant, and forPontedaria, one

110 B. Collins et al. / Ecological Engineering 23 (2004) 107–115Ta

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leaf per plant, was collected; the wet, underwater orfloating portion of the stem or leaf was bagged and thedry, above-water portion was discarded. For the plas-tic Juncusmimics, three stems per plant clump werecombined for one sample to approximate the same sur-face area of the real plant samples. Three plants perspecies per wetland were sampled. The plant sampleswere placed within a sterile stomacher bag and kept onice for transport to the laboratory.

In the laboratory, 20 ml sterile deionized water wasadded to each bag. The bags were resealed and pro-cessed in a stomacher lab blender (Teckmar) at normalsetting for 5 min. An additional 18 ml sterile deion-ized water was added to dilute the contents within thebag, and the contents were mixed by gently shaking.Twenty milliliter of the sample was pipetted into a ster-ile test tube. If this suspension contained a lot of algaeor chlorophyll, the sample was further diluted to re-duce color and avoid false positives. The sample wasthen vortexed and 150�l per well was inoculated intoa Biolog GN2 MicroPlateTM (hereafter Biolog plates).Biolog plates have 96 wells that contain 95 unique car-bon compounds, plus a control of distilled water. Whenbacteria utilize one of these carbon sources, tetrazoliumdye is reduced by bacterial respiration, and accumu-lates as insoluble formazin (Bochner and Savageau,1977). This results in the well turning from clear todarker shades of purple depending on the amount offormazin produced (Bochner, 1989; Garland and Mills,1991). The plates were sealed, incubated at 25◦C for7 r at5 asa l Sci-e herb ins

2

,1 dif-f ams( allp nt re-s ts> ea ess , the

2 h and read on a Biolog Microstation platereade90 nm. The intensity of the color in each well wnalyzed using Sigma Scan Pro software (Jandentific). The remaining plant pieces in the stomacag were dried at 50◦C for 24 h and weighed to obtaample dry weights.

.4. Data analyses

Analysis of dissimilarity (Clarke, 1993; Smith998) was used to determine if the Biolog scores

ered among the various treatments. Lance–WilliBray–Curtis) dissimilarity was calculated betweenlates. We tested the difference between treatmeponses (Clarke’sR: dissimilarity between treatmendissimilarity within treatments;Clarke, 1993; Clarknd Warwick, 1994). Because pairwise dissimilaritiharing an experimental unit are not independent

B. Collins et al. / Ecological Engineering 23 (2004) 107–115 111

probability of obtaining as large or largerRvalue wastested by 9999 permutations of treatment labels, strat-ified within each experimental block (row). The dis-tribution ofR values from the permutations was takenas the distribution ofR under the null hypothesis ofno treatment effect; significance was estimated as thenumber of permutationR values greater than or equalto the observedRvalue divided by (9999 + 1).

Differences in element (Fe, Mn, PO4, NH4, NO3)concentrations and pH in plants and wetland C efflu-ent were tested among planting treatments (real plants,artificial plants, no plants) and between aerobic wet-land order treatments (deep–shallow; shallow–deep) ateach sampling date using analysis of variance (whenvalues could be transformed to normality) or a non-parametric Kruskal–Wallis test. Differences betweentreatment pairs were tested by Bonferroni-adjustedt-tests or Wilcoxon tests. Rows were pooled when wet-land order was found to be nonsignificant.

3. Results

3.1. Influent and anaerobic wetland effluentchemistry

Between April 1999 and December 2000, in-fluent (into the CTW from the CPRB) was lowin pH and Ca (Table 1; Thomas, 2003). Acid-ity, which varied with rainfall, ranged from 738t ed1 tg on-c and7 overw lA w-i lA nd4 ons(

onw r-a t-l dity( d9 adn s in

the CPRB water and effluent from wetland A did notdiffer significantly throughout the experiment.

3.2. Aerobic wetland effectiveness

With treatments pooled, the aerobic wetlands hadlittle effect on pH and sulfate concentrations, but re-duced Fe, Al, and Mn concentrations 47, 80, and 63%,respectively (A wetland effluent compared to C efflu-ent,Table 1). Ca concentrations in wetland C effluentalso were 22% lower than those in A effluent. The pHof C effluent was within the range of that in uncon-taminated streams on the SRS (Table 1). Fe and Mnconcentrations in C effluent were slightly above, andAl and Zn concentrations slightly below, those in theuncontaminated streams (Table 1).

3.2.1. Planting treatment effectDuring the course of the experiment, significant al-

gal growth occurred in all aerobic wetlands, regardlessof planting treatment (real plants, artificial plants, noplants). Consequently, we estimated algal cover overthe wetland surface water to the nearest percent and an-alyzed chlorophylla andb concentrations in the algalmats. Although the algae were not analyzed or identi-fied in detail, it was noted that the algal mats differedamong the planting treatments. During summer, 1999,there was algal cover in 39–56% of the wetlands ineach row/planting treatment combination. Where algaewere present, algal cover ranged from 52± 39 to 82±2 df ionsrT aterq

erew ant-i lev-e hant -e hosei s inw gnif-i t fortl ntsw nts,t tions

o 2320 mg l−1 (CaCO3 equivalents) and averag304 mg l−1 (Table 1; Thomas, 2003). During the firsrowing season (April–October, 1999), Fe and Al centrations in the CTW influent averaged 1353 mg l−1, respectively. Concentrations decreasedinter (November 1999–March 2000; Fe = 114 mg−1,l = 59 mg l−1), then increased in the second gro

ng season (April–September 2000; Fe = 170 mg−1,l = 91 mg l−1). Mn concentrations averaged 3.7 a.2 mg l−1 in the first and second growing seasThomas, 2003).

Anaerobic wetland A had a significant effectater quality. Alkalinity in wetland A effluent aveged 619 mg l−1 (CaCO3 equivalents) and the we

ands neutralized >97% of the influent water aciTable 1; Thomas, 2003). Fe and Al were reduce0–99% (Table 1). However, the anaerobic wetland ho effect on Mn concentration; Mn concentration

9% and chlorophylla andb concentrations obtainerom the mats were at times quite high (concentratanged from 101.0± 59.6 to 166.6± 83.9�g l−1).hus, the effects of the planting treatments on wuality are confounded with algae.

Despite the confounding effects of the algae, there differences in water chemistry among the pl

ng treatments. For example, after May 1999, pHls were always lower in the tanks with real plants t

he other treatment tanks (Fig. 2). There were no differnces in pH between tanks with plastic plants and t

n the no plant treatment. Ammonium concentrationater passing through the real plant tanks were si

cantly reduced throughout the experiment excephe last sample collections (Fig. 2). Interestingly, NH4evels were low even during the winter when the plaere dormant. In the plastic and no plant treatme

here was a seasonal trend, with higher concentra

112 B. Collins et al. / Ecological Engineering 23 (2004) 107–115

Fig. 2. Mean monthly pH, NH4 (ppm), Fe (ppm), Mn (ppm), DOC(ppm), and PO4 (ppm), concentrations in effluent from aerobic wet-land C with real plants (P), plastic plants (F), or no plants (N). *indicates significantly (P < 0.05) lower concentration.

of NH4 during the growing season. Fe and Mn concen-trations also were affected by the presence of real plants(Fig. 2). Levels of these metals often were lower in theplastic plant and no plant treatments. There was a sug-gestion of a seasonal trend in the real plant treatments,with higher concentrations of these metals during thegrowing season.

Some measures of water chemistry did not differamong the planting treatments. Concentrations of dis-solved organic carbon (DOC) and phosphate showedsimilar trends in all wetland types, although eachshowed unique trends over time (Fig. 2). DOC concen-trations peaked in early summer, 1999, dropped in earlyfall, peaked again in late fall – early winter, and thenwere low throughout 2000. Phosphate concentrationswere higher during the growing season, and droppedover winter (Fig. 2). There were insufficient data to de-termine SO4 patterns among the planting treatments.

3.2.2. Wetland orderTank order in our wetland design did not affect wa-

ter quality. No significant differences between shallow-deep versus deep-shallow aerobic tanks were foundfor any metal or compound investigated (results notshown).

3.2.3. Bacterial community analysisBiolog plates determine the functional capacity of

bacterial communities to metabolize a variety of or-ganic compounds. Differences in processing as deter-m veld the-s epi-p esisi bac-t lastica sim-i ealp ia onp entf

realp be-t n thes nds,i em-b d (= no

ined by this method are indicative of community leifferences in the microbial assemblages. We hypoized that the geometry of the plants determines thehytic bacterial assemblage structure. If this hypoth

s true, there should be high similarity between theerial assemblages on real plants and those on pnalogs. In every comparison, there was higher dis

larity (P < 0.05) between than within plastic and rlant samples. Functional assemblages of bacterlastic analogs of real plants are significantly differ

rom those on their real counterparts.We next compared bacterial assemblages on

lants between different species within a wetland,ween species in deep and shallow wetlands, and oame plant species but in different order of wetla.e., deep-shallow or shallow-deep. Bacterial asslages between deep and shallow wetlands differeP0.003). Within the shallow wetlands, there were

B. Collins et al. / Ecological Engineering 23 (2004) 107–115 113

significant differences (P= 0.85) between bacterial as-semblages onJ. effususandP. cordatawhen stratifiedacross wetland order; however there was significantdifference between order of wetlands. Epiphytic bac-terial functions onJ. effususandP. cordatadiffereddepending on whether the wetland was second (rowB) or third (row C) in the three wetland system. Thebacterial assemblages onN. odorataandM. aquaticumwere significantly different from each other. However,epiphytic bacteria on these deep water plants did notdiffer depending on the order of the wetlands.

4. Discussion

Effluent from the passive treatment systems wassimilar in pH, higher in Fe and Mn, and lower in Aland Zn than surface water of uncontaminated streamson the Savannah River Site (SRS, 1998–2000). Treat-ment effectiveness of the systems compares favorablywith that of other wetlands designed to treat waste fromcoal or coal byproducts (e.g.,Brenner, 2001; Ye et al.,2001). The first wetland in the system, anaerobic wet-land A, increased pH and removed 90% of Fe and 99%of Al in the acidic, metal-rich influent from the CPRB,but had little effect on Mn concentrations. The aer-obic wetlands accumulated metals, including Fe, Al,Mn, and Zn, from the circumneutral water that hadpassed through the anaerobic wetland; with plantingtreatments pooled, the aerobic wetlands removed 47%o ndA thea se int 3%l RSs d unp tiono thefi ion,t ccu-m ass( g.,W al.,2 pol-i

de-s ofw d by

algae, which grew in 39–56% of the tanks in each plant-ing treatment and wetland order combination. Algalgrowth suggests metal concentrations in the circum-neutral water were not high enough to be toxic. Un-der nontoxic conditions, algae may take up metals inthe water column or leaked by plants (Jackson, 1998;Ye et al., 2001), and thus contribute to remediationof acid mine water (Kleinmann and Girts, 1987). Al-gae also utilize nitrates, sulfates, and phosphates, andcontribute to aerobic conditions in constructed treat-ment wetlands through photosynthesis (Gazea et al.,1996).

Although algal chlorophyll concentrations wereslightly higher in row B, order of the deep and shallowaerobic wetlands did not affect water quality. This sug-gests depth of the wetland and vegetation compositionat the inlet does not have an overriding effect on aero-bic wetland performance. Planting treatment, althoughconfounded by algae, affected some measures of waterchemistry, but not others. DOC and PO4 did not differamong the planting treatments. Water in wetlands withreal plants had lower pH and NH4, and higher Fe andMn, than water in those that had plastic plant analogs orwere unplanted. Plants utilize N and can contribute tolower pH through respiration and litter decompositionprocesses. Lower pH in the wetlands with real plantsmay have contributed to their higher Fe and Mn con-centrations. Precipitation of both these elements maybe inversely related to pH, and Mn removal through pre-cipitation may only occur when pH is≥ 7 (Wildemane 01;Y ira mayh ccu-m ectso

meo ught nala withr ffer-e shal-l es),ad canb itiesd thec ytic

f the Fe and 80% of the Al in the influent from wetla. Although Zn concentrations in the effluent fromerobic wetlands averaged 1.8% greater than tho

he influent from anaerobic wetland A, they were 7ower than Zn concentrations in uncontaminated Streams. These results suggest both the planted anlanted aerobic wetlands can contribute to remediaf acidic drainage from coal processing facilities asnal part of a passive treatment system. In addithe plants in deep and shallow aerobic wetlands aulated a minor fraction of P and metals in biom

Collins et al., 2004), which supports conclusions (e.ildeman et al., 1993; Gazea et al., 1996; Ye et

001) that planted aerobic wetlands can serve asshing components in treatment systems.

The aerobic wetland treatments, which wereigned to identify the most effective combinationetland depth and composition, were confounde

-

t al., 1993; Sparling and Lowe, 1997; Brenner, 20e et al., 2001; Lee et al., 2002). Thus, despite thebility to accumulate elements in biomass, plantsave an undesirable effect on water quality when aulation does not balance or exceed lower pH effn metal solubility.

Although the relationship is not clear, at least sof the effect of plants on water quality may be thro

heir effect on the microbial community. Functiossemblages of bacteria differed between tankseal plants and those with plastic plant analogs. Dinces in microbial assemblages between deep and

ow wetlands (which contained different plant specind betweenN. odorata and M. aquaticum in theeep tanks, suggest plant–microbial interactionse species-specific. Because microbial communiffered between species of plants, differences inhemistry of tanks may reflect differences in epiph

114 B. Collins et al. / Ecological Engineering 23 (2004) 107–115

microbial processes. Further research is needed to ad-dress this specificity.

We conclude that plants do have an effect on wa-ter quality, in part because they affect bacterial assem-blages. Planted aerobic wetlands can help remediateacidic, metal contaminated water through plant uptake,but may contribute to lower pH and greater levels of Feand Mn in treatment wetland effluent if uptake doesnot exceed pH effects on solubility. Further researchis needed to address the specificity of plant–microbialinteractions, which suggest that wetland design deci-sions, including wetland depth and species planted,could exacerbate or reduce plant effects on waterquality.

Acknowledgements

Robert Thomas and Daniel Coughlin were thetechnical and motivational forces for this research,which was funded by the National Water ResearchCenter as part of Financial Assistance Award numberDE-FC09-96SR18546 between the U.S. Departmentof Energy and the University of Georgia’s SavannahRiver Ecology Laboratory.

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